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    Chasing crayfish and the leeches that live on them

    “I study leeches called branchiobdellidans that live on crayfish. These leeches are just millimetres long and are symbionts — meaning they live in a close, long-term association with their host. They feed on microorganisms and debris that collect on the host’s surface. In small numbers, they help to keep the crayfish clean, but in large colonies, they can become mildly parasitic.Here, I’ve just caught a crayfish under torchlight, and I’m holding it carefully so that the symbionts aren’t washed away. After catching each crayfish, I measure its length and use a paintbrush to push a symbiont sample into a small vial. Tweezers would damage their delicate bodies.This photo was taken in June, in a small river in Slovenia where I recently discovered a new species of symbiont. Later, in the laboratory, I analysed the sample’s DNA to trace how Astacus astacus — the noble crayfish — and its symbionts have evolved together over millions of years.The leeches aren’t just passengers; they’re bioindicators. When they disappear, it can be a signal that crayfish populations — and the rivers themselves — are in trouble. Invasive crayfish from North America (Pacifastacus leniusculus), which were introduced for farming, are already disturbing the ecological balance in waterways in Slovenia.

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    Microplastic removal across ten drinking water treatment facilities and distribution systems

    AbstractThe performance of conventional and advanced drinking water treatment processes for the removal of microplastics is poorly understood due to the use of a wide range of methods for sample collection, isolation, and analysis that make direct comparison among studies challenging. In this study, microplastic (>2 µm) removal across ten drinking water treatment facilities, as well as their presence in source waters and distribution systems, was characterized. Municipal drinking water treatment facilities achieved >97.5% removal, primarily due to chemically assisted granular media filtration or ultrafiltration. In untreated source waters, concentrations ranged from 1193 ± 64 to 7185 ± 64 particles/L, with polypropylene, polyethylene, polyamide, and plastic copolymers representing the most common polymer types identified. These findings provide insight regarding microplastic exposure via drinking water, as well as treatment process performance for their removal which may be used to inform the development and implementation of future regulations and/or guidelines.

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    IntroductionMicroplastics (MPs) have been reported in both source and drinking waters around the globe1,2,3,4,5,6,7,8,9,10,11,12,13,14,15,16,17,18. Drinking water serves as a direct exposure route along with food, air, and skin contact19, highlighting the importance of their quantification as well as removal during treatment. However, studies which focus on drinking water have employed a wide range of methods for collection and isolation of MP particles, making it challenging to directly compare results regarding treatment process performance across different facilities. Monitoring of MPs in drinking water has been legislated in the States of California (Senate Bill 1422, 2018) and New Jersey (Assembly Bill 4821, 2023), further emphasizing the need for a robust and practical assessment methodology to be developed and widely adopted20,21. The ability to capture and quantify MPs <30 µm is of particular importance due to potential implications to human health22,23,24,25. Many methodological challenges exist when sampling MPs in drinking water. For example, deposition of airborne MPs may potentially contaminate samples and should be mitigated by employing enclosed in-line sampling methods26 rather than use of bottles1,2,8,9,16,17. No single study has systematically investigated the occurrence and removal of MPs <30 µm from a wide range of treatment processes in order to obtain a representative estimate of exposure via drinking water.In this study, an extensive MP sampling campaign included ten conventional (employing chemically assisted granular media filtration), as well as advanced drinking water treatment facilities (employing micro- or ultrafiltration). MP quantification, size measurement, and chemical identification of particles ≥2 µm were conducted using microscopy coupled with Raman spectroscopy. Facilities were selected to represent: (1) a range of processes used to treat water originating from the same source, and (2) similar treatment processes used to treat different source waters. Study sites were geographically diverse with source waters encompassing four lakes and two rivers (Table 1). Selected facilities served by the same source waters were sampled on different days, thus providing an indication of potential MP temporal variability. Treatment processes included those classified as conventional (A, B, C, D, E, and F), as well as advanced which included ultrafiltration (UF) (G, H, and I) or microfiltration (MF) (J). Since the primary purpose of filtration is to remove particulate matter, it is anticipated to contribute substantially to MP reduction, as previously reported in studies regarding conventional13 and advanced27 treatment processes. Coagulation, flocculation, and sedimentation (CFS) have been suggested to serve as primary contributors to MP removal rather than filtration1,9,15,28,29. As such, samples were collected following both CFS and filtration, wherever possible. Other treatment processes included biological activated carbon (BAC), ozone, and ultraviolet (UV) disinfection (Tables S1 and S2). Microorganisms present in BAC contactors have been hypothesized to bioregenerate adsorption sites by consuming adsorbed organic compounds, thus enabling biodegradation and adsorption to continue over extended periods of time30, which may also contribute to MP removal31. Previous studies have suggested that ozone may cause fragmentation of MPs, resulting in an increase in concentration1,15. Collection of samples prior to and following ozonation was conducted to specifically address this question. In contrast, disinfection using UV or UV-based advanced oxidation are not anticipated to result in MP fragmentation: concentrations of surface functional groups of low- and high-density polyethylene as markers of oxidation, were not observed to be impacted when applying UV or UV/H2O2 at dosages typically employed in water treatment32, although leaching of organic matter may be possible33. To examine the performance of specific treatment processes, samples were collected from: (i) untreated source water, (ii) following each major unit treatment process, (iii) finished water (following disinfection), and (iv) the nearest regulatory sampling point in the distribution system (hereafter referred to as “distribution”). To capture MPs, water was filtered onsite using an enclosed, stainless-steel inline system which contained 10 µm and 2 µm polycarbonate (PC) filters arranged in-series34. Prior to analysis using Raman microscopy, samples were subjected to Fenton oxidation, followed by enzymatic digestion using cellulase and trypsin (method adapted from Cheng et al.35) to reduce the presence of non-plastic particulates.Table 1 Sampling locations and associated treatment process informationFull size tableMicroplastic concentrations in untreated source watersMP concentrations in source waters ranged from 1193 ± 64 MP/L to 7185 ± 64 MP/L (Fig. 1, Table S3). Previous studies which considered a minimum particle size of 1 µm have reported concentrations within the same order of magnitude (multiple thousands of MP particles per liter)1,8. Concentrations for two treatment facilities which shared the same intake and were collected on separate days (Lake 3, F and H) ranged from 2521 ± 64 MP/L to 4574 ± 64 MP/L. A comparable 2668 ± 64 MP/L was observed for the influent of facility E (Lake 3 at an intake 50 km away from F and H). Varying MP concentrations for the same source water at different times suggest potential temporal and spatial variability. Repeated monitoring over time is required to investigate temporal changes more thoroughly.2 µm) at ten conventional and advanced drinking water treatment facilities.”>Fig. 1: Comparison of estimated microplastic concentrations (>2 µm) at ten conventional and advanced drinking water treatment facilities.Conventional treatment: facilities that utilize granular media filtration (A/S anthracite over sand, GAC/S granular activated carbon over sand). Advanced treatment: facilities that utilize ultrafiltration (UF) or microfiltration (MF). Vertical bars represent the propagated error following blank subtraction (n = 3). ND non-detectable following blank subtraction.Full size imageA Pearson correlation analysis indicated a weak, non-significant negative relationship between intake depth, which ranged from 3 to 26 m, and source water MP concentration (r = −0.213, p = 0.555). Although intake depth may impact MP concentration, the presence of MP particles in a given water column may also be influenced by polymer density, morphology and size, as well as biofouling36,37. Stratification, turbulence, and sediment resuspension can redistribute MPs37,38,39 and reduce depth-related patterns37,40,41,42. As such, intake depth alone cannot serve as a reliable indicator of MP concentration.While a surrounding population (Table 1) is often considered to influence MP concentrations at intakes, several other factors may include land use43,44, proximity to point sources (e.g., wastewater discharges)45, and hydrological conditions (e.g., flow velocities, retention of MPs in soils, vegetation, or upstream reservoirs)37,44,46,47. Several of the source waters included in this study share overlapping watersheds, making it challenging to identify the influence of a specific population center.MP concentrations in the lakes exhibited greater variability as well as higher maximum values when compared to rivers (Fig. S1). Lakes may act as sinks for MPs48, but the extent is influenced by surrounding land use, direct point-source pollution, inflow characteristics, and hydrodynamics43,44. The facilities considered in this study primarily received water from large lakes where dilution may serve to reduce MP concentrations44. As a result of these complex factors, MP quantification must be conducted on a site-specific basis.Microplastic removal in drinking water treatment facilitiesFinished water concentrationsWhen considering all ten facilities, MP concentrations in finished waters ranged from non-detectable (facilities G and H) to 64 ± 64 MP/L (facility C) (Fig. 1, Table S3). Following application of a one-way ANOVA with post-hoc Tukey’s HSD, a significant reduction in MP concentrations was observed when comparing source waters to finished waters (F = 29.1, p = 1.82 × 10⁻7; Fig. S2), confirming effective removal during treatment. Finished water concentrations in facilities employing UF (G, H, and I) significantly differed from those employing MF (J) or granular media filtration (A-F) (F = 7.94, p = 0.0226, Fig. S3). Granular media is capable of removing particles 0.1 to 1 µm49, similar to MF with a pore size of 0.1 µm (Table S2). Lower MP concentrations (ranging from non-detect to 12.7 ± 64 MP/L) were observed following UF with a pore size of 0.02 µm (Table S2) when compared to post-filtration at facilities that employed granular media filtration or MF (9.3 ± 64 to 400.2 ± 64 MP/L). Theoretically, particles >2 µm should not be present following membrane filtration where nominal pore sizes range from 0.01 to 0.2 µm50. However, samples could not be collected directly from membrane permeate, but rather at the first accessible location immediately downstream, typically a clearwell, which can be susceptible to airborne MP deposition15.Total microplastic removalTotal MP removal was calculated by comparing concentrations in finished waters to those in source waters (Table S3). For facilities B and G, removal was based on the first sample collected in the treatment process (i.e., post-sedimentation for facility B and post-ozone for facility G), as source water data was not available. When directly comparing conventional facilities (A–F), removals ranged from 98.7% to 99.4%. For the one facility that employed MF (J), removal was 97.5%, whereas the three facilities that incorporated UF (G, H, and I) achieved 99.3% to 100%. A significant difference in removal (one-way ANOVA, F = 6.39, p = 0.0354; Figure S4) was observed when comparing UF facilities (G, H, and I) to MF (J) as well as to conventional facilities (A-F). This difference may be attributed to the smaller UF pore size (0.02 µm). A previous bench-scale study involving similar UF membranes reported complete rejection of MPs > 1 µm27. Removal observed across all facilities was >97.5% despite the significantly greater removal for those which employ UF (Fig. 1). Previous studies have reported treated water concentrations (particles ≥ 1 µm) to range from 18.7 to 1401 MP/L1,2,8,17,51 with corresponding removals of 50% to 88%. Higher removal efficiencies observed in the current study (97.5% to 100% for MPs ≥2 µm) can be attributed to differences in minimum particle size thresholds, sampling and analytical methods, as well as operational parameters. Several previous studies quantified particles (>1 µm) using scanning electron microscopy1,8,9,51 and subsequently characterized polymer types based on a subset of larger particles (≥10 µm8 or ≥50 µm in size51) using Raman or FTIR spectroscopy29,42. In the present study, MP quantification, size, and chemical identification of particles ≥2 µm were all determined via microscopy coupled with Raman spectroscopy.Removal by individual treatment processesFiltration, either via granular media or membranes (MF or UF), provided the greatest contribution to removal, ranging from 87.5% (facility B) to 100% (facility G) (Table S3). Statistical analysis could not be conducted with respect to MP removal via individual treatment processes or parameters such as granular media filter depth, membrane age, sedimentation, ozonation, or BAC filtration (Tables S1 and S2) due to variability in treatment design and operation. When considering conventional facilities, qualitative assessment suggested that specific media type (e.g., granular activated carbon vs. anthracite over sand), filter loading rate, bed depth, or effective grain size did not have an impact on MP removal. For advanced facilities, the age of UF or MF membranes did not appear to impact MP removal.While several previous bench-28,52,53,54 and full-scale1,9,55 studies reported coagulation-flocculation-sedimentation (CFS) to provide MP reduction, this was not apparent in the current study (Fig. 1). Studies by others also suggested that CFS does not provide a barrier for small MPs (i.e., <20 µm), as they are less effectively removed by Al- and Fe-based coagulants when compared to particles ≥20 µm54. In this study, facilities D, F, and I exhibited increased MP concentrations post-sedimentation (compared to source water), whereas facilities A and C exhibited a decrease (Table S3). A shift in particle size distribution towards smaller particles following CFS was not observed, suggesting that it did not preferentially remove larger particles. In addition, no consistent MP removal trends were observed with respect to BAC age, or the application of ozone or UV.Impact of particle morphology, size, and polymer type on removalWastewater effluent, airborne MPs, and surface runoff are typically believed to contribute MP fibers56,57,58,59, defined as particles with a length-to-width ratio >360. As such, source waters may be anticipated to contain a larger proportion of fibers (versus fragments), reported by others to comprise 37% to 61%8 and 54% to 74%1 of total MP concentrations. In contrast, among the ten facilities examined in this study, fibers comprised 7.8% and 7.5% of all MPs in source and finished waters, respectively (Table S3). When considering total MPs in all samples, only 6.9% were fibers, with fragments being the predominant morphology. The relatively low observed proportion of fibers may be due to potential fragmentation during Fenton oxidation (especially PET and PA fibers which contain ester and amide bonds, respectively) and/or digestion by cellulase (for cellulosic fibers). In agreement with previous studies1,8, the relative proportions of fragments to fibers did not change when considering a wide range of source and finished waters, suggesting that morphology did not impact removal during treatment.When considering size, defined as the major dimension of a given MP particle, 81.1% ± 9.4% of all MPs were <20 µm; 50.6% ± 10.4% were <10 µm (Fig. 2, Table S6). As previously reported, removal of MPs during drinking water treatment is anticipated to increase as a function of particle size1,8,9,61. In this study, a reduction in the proportion of MPs >20 µm was observed for facilities A, B, C, F, and J, whereas for D, E, G, H, and I they remained the same or increased (Table S6). When considering all finished waters, MPs <20 µm comprised >75% of all MPs observed, suggesting that previous studies (which only characterized particles > 20 µm) may have unintentionally excluded a substantial proportion of MPs present in drinking water. Thus, it is recommended that future studies employ Raman microscopy for polymer type identification, quantification, and measurement of MPs >1 µm. Accurate reporting of MP count, size, and polymer type is especially important given that particles <30 µm may translocate within humans25. The relative abundance of particles in the 2–5 µm size fraction is lower than anticipated based on the overall decreasing trend with increasing particle size (Fig. 2). This discrepancy likely reflected the narrower bin width (3 µm vs. 5 µm in other bins).Fig. 2: Relative size contribution of averaged source, finished and distribution system samples (all facilities).Boxes represent the inter-quartile range (IQR). Horizontal lines represent median values. Vertical lines represent values within 1.5× the IQR. Points represent outliers beyond 1.5× the IQR. Size is defined as the major particle dimension.Full size imagePolymer type was dominated by polypropylene (PP), polyethylene (PE), polyamide (PA), and copolymers in all samples (Tables S4 and S5), with PP typically being the most abundant (average relative contribution 61.5% ± 23.3%). Values were below detection limits for facility G. Previous studies have reported PP and PE to be the most abundant types present in source and treated waters1,7,8,9,61,62,63,64,65. Source waters associated with facilities G and I exhibited a large proportion of polyester (31.1% and 25.0%, respectively) when compared to others (1.9% ± 2.5%). This discrepancy may be due to environmental and temporal variability since facilities F and H share the same intake as G but were sampled at different dates. Seasonal flow dynamics, storm events, or variations in MP input to source waters could influence the types and proportions of MPs present at any given time47,48. As such, the absence of elevated polyester levels at facilities F and H compared to facility G may reflect short-term or seasonal variability.The impact of polymer type on removal was qualitatively examined (Table S4) for individual treatment facilities by considering both source and finished waters (Fig. 3). Neither conventional nor advanced treatment processes preferentially removed any specific polymer type, implying that removal mechanisms may be similar to those for inorganic particulates and not impacted by chemical composition, as reported by others1,9. Filtration (by granular media or membrane filtration) removes particles based on physical exclusion or adsorption66. This functional similarity likely contributed to the lack of preferential removal of specific polymer types as observed in this study.Fig. 3: Relative contribution by polymer type in source, finished, and distribution water for both conventional (A, B, C, D, E, and F) and advanced treatment facilities (G, H, I, and J).Copolymer refers to a combination of two or more polymer types. ND: non-detectable following blank subtraction. PP polypropylene, PE polyethylene, PA polyamide, PEST polyester, PDMS polydimethylsiloxane, PS polystyrene, PVC polyvinyl chloride.Full size imageMicroplastics in distribution systemsNo statistically significant difference in MP concentrations was observed when comparing finished water and distribution system water (Tukey’s HSD pairwise comparisonFW vs DS, p = 1.00, Fig. S2). It is hypothesized that deposition of airborne MPs into reservoirs may potentially represent a source of MPs within distribution systems15 but was not observed in this study. In addition, relative polymer type abundance and size distribution were similar when qualitatively comparing finished and distribution system water (Figs. 2 and 3, Tables S5 and S6). It should be noted that hydraulic residence time was not considered; sampling locations represented those employed for regulatory monitoring purposes and in general were supplied by large diameter (≥900 mm) concrete pressure pipes.The distance traveled through a given distribution system was not observed to have a significant impact on MP concentration when comparing finished and distribution system water (Pearson correlation, r = −0.202, p = 0.5752; Fig. S5). It is suggested that more extensive sampling throughout distribution systems be conducted in future studies to assess the influence of pipe material and age, hydraulic residence time, flushing/cleaning practices, as well as potential exposure to airborne MPs in reservoirs.Correlation with surrogate water quality parametersRoutinely monitored water quality parameters including turbidity, total particle counts, UV254, specific UV absorbance at 254 nm (SUVA), pH, total and dissolved organic carbon (TOC and DOC), were measured for all samples (Table S9). Specific fractions of natural organic matter (NOM) in source and finished water were also quantified (Table S10). Only parameters that represent particulate material (turbidity, total particle counts, and TOC) were examined for potential correlation with MP concentration. Prior to analysis, UV254, SUVA, DOC, and NOM samples were passed through 0.45 µm filters and as such represent dissolved material.A Spearman correlation analysis using source water values revealed no significant correlations between MP concentration and turbidity (r = −0.249, p = 0.487) or total particle counts (r = −0.297, p = 0.405) (Fig. S6). A significant negative correlation was observed between MP concentration and TOC (r = −0.758, p = 0.0111) (Fig. S6), which may indicate sites with higher natural organic matter to be less impacted by anthropogenic MP inputs. However, this relationship should be interpreted with caution, as the apparent correlation may be influenced by other site-specific conditions including organic loading, hydrodynamics, and proximity to MP inputs. The absence of a significant relationship between MP concentration and turbidity is likely due to the fact that source waters also contain many particulates of which microplastics only represent a small fraction.When considering finished water, no significant correlation was observed between MP concentration and turbidity (r = −0.483, p = 0.157), TOC (r = −0.288, p = 0.420), or total particle counts (r = 0.122, p = 0.738) (Fig. S7), again indicating that traditional water quality parameters do not correlate with MP concentrations. Similarly, total MP removal did not correlate significantly with the removal of either turbidity (r = −0.0791, p = 0.828), TOC (r = −0.289, p = 0.418), or total particle counts (r = 0.0669, p = 0.854) (Fig. S8). These findings emphasize that it is essential to implement a dedicated monitoring program in order to quantity MP concentrations, specific polymer types, and sizes.ConclusionsMicroplastics were consistently present in source waters including lakes and rivers with considerable variability in concentration and polymer composition. Most MPs consisted of fragments <20 µm, predominantly represented by PP, PE, PA, and copolymers. Drinking water treatment facilities which employ a range of conventional and advanced processes were capable of removing >97.5% of MPs ≥2 µm. Chemically assisted granular media filtration or membrane filtration provided the highest removal, while coagulation, flocculation, and sedimentation, ozonation, and UV disinfection were observed to have minimal impact. The similarity in removal across a range of particle sizes, shapes, and polymer types suggests that physical exclusion, rather than chemical interaction, is predominant. A lack of correlation between MP concentration and conventional water quality parameters, such as turbidity or TOC emphasizes the need for direct measurement using dedicated analytical methods. Overall, these findings demonstrate the capability of drinking water treatment practices in reducing MP exposure to consumers, while highlighting the necessity for standardized protocols to ensure accurate assessment. Future studies should further explore the impacts of specific distribution system factors such as pipe type, age, water velocity, and hydraulic residence time.MethodsSampling locationsMPs were collected from ten drinking water treatment facilities located across southern Ontario, Canada, representing a range of source waters and treatment processes (Table 1). These facilities span a range of urban settings and are situated in areas influenced by different watershed characteristics and climatic conditions typical of the Great Lakes Basin, including temperate seasonal variability with cold winters and warm, humid summers. At all facilities, single samples were collected at the intake (defined as “source water”), immediately following each individual treatment process (whenever possible), as well as finished water (after the final stage of treatment). Distribution system samples were collected at the first downstream sampling location typically used for regulatory compliance monitoring. Conventional treatment facilities are classified as those which employ granular media filtration, whereas advanced treatment facilities incorporated micro- or ultrafiltration. Coagulation and flocculation were practiced at all facilities with the exception of E, G, H, and J. Samples were collected from the effluent of individual filtration processes for facilities that employed two different types of granular media filters operated in parallel (anthracite over sand and granular activated carbon (GAC) over sand, facilities A and E, respectively), or biologically activated carbon (BAC) (facility J). Process-specific characteristics which included grain size (granular media filtration), ultra- or microfiltration membrane age, and UV and chlorine dosages are shown in Tables S1 and S2.Microplastic sampling equipmentDue to the extensive scope of this study, which involved sampling ten different drinking water treatment facilities across southern Ontario, only one sample per treatment step per facility was collected. This approach was necessary to balance the practical constraints of fieldwork logistics, including the considerable travel distances between facilities and the time-intensive nature of MP sampling and analysis. The sampling strategy prioritized capturing a broad representation of different treatment technologies and operational conditions across the region, in lieu of repeated sampling at individual facilities. While this potentially limits the ability to assess short-term variability associated with source water intakes and within treatment processes at a single site, it provides valuable insight regarding the overall pattern of MP removal across a diverse range of full-scale drinking water treatment operations.To collect MPs, an enclosed in-line filtration apparatus consisting of two stainless-steel, 47 mm diameter filter holders (MilliporeSigma, Darmstadt, Germany) connected in-series was employed as described by D’Ascanio et al.34. Particulates were isolated onto 10 and 2 µm, 47 mm diameter polycarbonate (PC) filters (MilliporeSigma, Darmstadt, Germany) enclosed within the holders. Quick disconnect brass couplings and control valves were located at both the inlet and outlet. Threaded connections were sealed using polytetrafluoroethylene (PTFE) tape. Rubber O-rings were used to seal the interior of the filter holders, and to prevent potential contamination by airborne MPs. Only facility E was sampled using only a 5 µm PC filter (Sterlitech, Auburn, Washington, United States) upstream of a 2 µm PC filter (instead of a 10 µm) filter due to rapid occlusion of the 2 µm filter by particulates during initial sampling trials. Specific polymer types associated with sampling or lab equipment were excluded from analysis, including PTFE, PC, nitrile glove material and rubber O-rings. All “MP-free” reagents (Table S12) were filtered through a 0.45 µm PTFE filter (Omnipore, Millipore Sigma, Darmstadt, Germany) and stored in rinsed, closed glass bottles.Prior to sampling, the filter holders were sonicated in MP-free water for 30 min and subsequently cleaned using MP-free water which consisted of 0.05% Tween® prepared with ultrapure water (18.2 MΩ·cm, 0.2 µm filtered, Milli-Q EQ 7000 System; MilliporeSigma, Darmstadt, Germany). All cleaning and assembly were conducted within a Class II laminar flow hood to minimize the potential for contamination by airborne particles. All glassware and equipment including glass filtration apparatuses, Pasteur pipettes, 600 mL glass beakers, 1 and 4 L glass bottles, PTFE stir bars, sampling equipment, and stainless-steel forceps were rinsed three times using MP-free water. Following placement of PC filters inside the in-line filtration holders, brass control valves at each end (inlet and outlet) were closed and covered with aluminum foil to minimize any potential contamination by airborne MPs.At all water treatment facilities, individual samples were collected starting with finished water and ending with source water. Samples were collected directly from the influent or effluent of specific treatment processes via sampling taps, whenever possible. When sampling line pressure was ≥30 PSI, in-line filter holders were connected directly. For pressures <30 PSI, or when an air gap was requested by facility personnel, a rotary vane pump (GA072, Fluid-o-Tech) and a 30 L stainless steel constant head tank (CHT) equipped with an overflow port at 23 cm, inlet port at 17 cm, and outlet at 4.5 cm from the bottom was employed (Fig. S9). The tank was covered with a stainless-steel lid. All tubing upstream of the in-line filtration apparatus to the sampling port consisted of braided stainless steel. A pressure relief valve set to 50 PSI (Apollo Valves, Matthews, North Carolina, United States) was incorporated upstream of the filter holders to ensure PC filter integrity.Prior to sample collection, all components (including the pump and CHT when required) were initially flushed for 15 min following connection to the sampling port. The in-line filter was then connected to collect a pre-selected volume of water. Due to higher turbidities upstream of granular media filtration or UF, lower sample volumes (typically 0.5 L) were collected when compared to post-filtration (typically 50 L). Sample volumes initially collected from facility G were used to inform those subsequently collected for both pre- and post-filtration at other facilities. Particle concentrations per analyzed filter area obtained from facility H samples were used to inform pre-filtration sample volumes that would result in a number of particles that could be analyzed within 2–3 days based on the Raman microscopy procedure used in this study (Table S3). Volumes ≤ 0.5 L were measured using a 1 L graduated cylinder placed at the discharge end of the in-line filter, whereas volumes > 0.5 L were quantified using a positive displacement low flow water meter (Assured Automation, Roselle, New Jersey, United States) at the outlet of the in-line filtration apparatus. Following collection of a desired volume, inlet and outlet control valves were closed and the in-line filter assembly disconnected. Inlet and outlets were covered with aluminum foil for transport to the laboratory.Turbidity and pH were measured in the field using handheld instruments (HF Scientific MicroTPW Portable White Light Turbidimeter and Orion Star A121 Portable pH Meter, Fisher Scientific, Ontario, Canada). Parallel samples were also collected for in-lab analyses which included TOC, DOC, UV254, specific UV absorbance (SUVA), particle count, turbidity, and pH. Additional samples of raw and treated water were collected at each facility (except for G) for liquid chromatography organic carbon (LC-OCD) analyses. Complete water quality data is shown in Table S9; LC-OCD results are shown in Table S10.Analytical methodsExtraction and digestionA vacuum pump was used to filter any residual liquid in the filter holders (inside a Class II laminar flow hood). The in-line filter holders were then disassembled and the two PC filters placed into 600 mL beakers. All interior parts of in-line filters (including stainless steel underdrains and support screens), were rinsed three times into the same 600 mL glass beakers using a minimum of 5 mL of MP-free water. PC filters and associated water were sonicated for 55 min at 25 °C. PC filters were removed using forceps and rinsed into a beaker using MP-free water. The suspension was filtered through a 1 µm PTFE filter (Omnipore, MilliporeSigma, Darmstadt, Germany) using vacuum filtration. The PTFE filter was then subjected to Fenton oxidation, followed by cellulase and trypsin enzymatic digestion following a method adapted from Cheng et al.35. Reagent specifications and operational parameters are provided in Table S14.Following digestion, each sample was filtered using a 1 µm, 47 mm diameter PTFE filter for analysis. Prior to analysis using Raman microscopy, PTFE filters were mounted on a custom built stainless- steel stage, specifically designed to ensure that the filters were held taut to provide a flat surface for analysis.Raman spectroscopy methodologyRaman spectroscopy coupled with microscopy was performed using a Horiba XPlora Plus system with LabSpec 6.6 software (Version 6.6.2.7, Horiba Scientific, Kyoto, Japan). Instrument parameters are provided in Table S15. A spectral range of 150–3600 cm-1 was employed to include the entire range where peaks characteristic of plastics are known to occur67. The laser power at the surface of the filter was 7.54 mW at 100% when using a long working distance (LWD) 50× objective (LMPlanFL N 50x/0.50 BD, Olympus, Richmond Hill, Ontario, Canada). The system was calibrated every 24 h to a 520.7 cm−1 peak obtained from a silicon wafer.Evaluation of the entire area of a 47 mm diameter filter would result in unrealistic analysis times, exceeding two weeks for a single sample. Initially, 6.6% of the filter area was analyzed (facilities G and H), as adapted from Pittroff et al.10. The filter area was divided into two zones: an inner circle with a diameter of 25 mm and an outer ring with an inner diameter of 25 mm and an outer diameter of 35 mm; both zones had equal surface areas (Fig. 4). Particles in these zones were analyzed to account for potential variations in particle concentration between the center and edge of the filter. For all other facilities, 0.45% to 5.42% of the total area was analyzed to ensure that a minimum of 40,000 to 50,000 individual particles were considered, which typically could be accomplished over two to three days. For each filter area, a mosaic image was obtained using the ViewSharp tool to create a corresponding topographic map using the 50× LWD objective with bright field illumination. Mosaic images were then analyzed using the Particle Finder tool to identify individual particles and subsequently collect corresponding Raman spectra.Fig. 4: Locations of four subsections analyzed within two zones on filter surfaces.The inner shaded circle and outer ring represent equal surface areas.Full size imageA two-phase spectral acquisition approach was applied to optimize overall analysis time. During Phase I, spectra were acquired, baseline corrected, and screened for the presence of MP particles based on the C-H stretching region observed for most polymer types (except PTFE) from 2800 to 3150 cm−1 10,68,69 by employing the parameters and settings described in Table S13. In Phase II, post-acquisition and correction screening was applied to identify potential plastic polymers. Following screening, points on the mosaic corresponding to peak areas >2 were selected for further analysis where higher quality spectra were acquired for subsequent comparison to libraries of known polymers. Following baseline-correction, spectra were imported using WITec TrueMatch (6.1.7, WITec Wissenschaftliche Instrumente und Technologie, Ulm, Germany) for identification using three databases: (i) S.T. Japan (S.T. Japan Europe GmbH, Köln, Germany), (ii) SLOPP, and (iii) SLOPP-E6, as well as an in-house database associated with 21 polymers from Polymer Kit 1.0 (Hawai’i Pacific University Centre for Marine Debris Research). In order to reduce any potential influence of the PTFE filter on HQI values of suspected polymers, the region of 709–759 cm−1 3,4 was excluded from the database search. To reduce potential overestimation of MPs, those identified as polymers matching items used in sample collection/extraction (e.g., rubber O-rings inside the in-line filter holders, nitrile gloves) were disregarded. All spectra were manually confirmed; material types were sorted into groups using a method adapted from Munno et al.70 (Table S5). As described in Table S5, anthropogenic cellulose and anthropogenic unknown particles were identified, but excluded from further calculations. Anthropogenic cellulose was defined as particles with a spectrum consistent with dyed cotton or dyed cellulose, whereas the anthropogenic unknown class unites all additives, dyes, pigments and material which indicate anthropogenic origin, or modification without definitive plastic or cotton/cellulose indications.QA/QC proceduresContamination control measuresWhenever possible, all laboratory work was conducted in a Class II laminar flow hood. White 100% cotton lab coats and nitrile gloves were worn at all times by lab personnel. All glassware and equipment used for sampling, cleaning, digestion, and spectroscopic analysis was rinsed three times with MP-free water. Areas in the lab where MP sample digestion or analysis were conducted (outside of a Class II laminar flow hood) were equipped with portable high efficiency particulate air (HEPA) filtration systems (Coway, Las Vegas, Nevada, United States) to reduce any potential airborne particulates. Throughout all analyses, samples and MP-free water were kept covered whenever possible.Laboratory blanksThree laboratory blanks were prepared by placing new 10 µm and 2 µm, 47 mm diameter PC filters within a cleaned in-line filtration apparatus, removing it from the Class II laminar flow hood (to simulate transporting equipment to the field for sampling), and then returning it to the laminar flow hood for subsequent MP extraction. The same procedures for the extraction and digestion of actual samples were employed for laboratory blanks. To address any potential changes in lab or operating conditions, blanks were conducted at time intervals evenly spread across the entire study period, typically one blank per 25 samples with a laboratory blank approximately every 4 weeks.Data processingMPs identified in analyzed filter areas were extrapolated to the entire filtration area (979 mm2). Mean extrapolated MPs measured in blanks were subtracted from the extrapolated MPs within actual samples. Particle concentrations (MP/L) were calculated by dividing blank-subtracted values by the volume of water filtered. Standard deviation of the mean number of MPs for blanks (n = 3) was used to indicate the variation associated with blank subtraction.Statistical analysesQuantitative analyses were carried out wherever data availability and consistency permitted. In cases where a dataset was too small or exhibited high variability in operational parameters (e.g., membrane age, sedimentation and filtration parameters), qualitative assessments were conducted to support interpretation.Statistical analyses were conducted using Python. Normality of the data was assessed using a Shapiro-Wilk test. Results can be found in Table S11. Depending on the distribution of the data, either Pearson or Spearman correlation coefficients were calculated.Group differences were evaluated using one-way ANOVA, followed by Tukey’s HSD post-hoc testing where applicable. Results are shown in Figs. S1–S8.Study limitationsReported concentration values (MP/L) represent extrapolations based on the analysis of 0.45–5.42% of total filter area. Only MPs ≥ 2 μm were considered due to the minimum pore size of the filters employed during sampling. Given that only a percentage of each sample was analyzed via Raman spectroscopy, the resulting sub-sampling introduces a deviation of approximately 17% or more71. Thus, reported values should be interpreted as estimates rather than precise quantifications.Raman spectra were acquired using a method that may exclude particles which exhibit fluorescence or degrade via heating and/or photodecomposition resulting in a peak area of <2 within the 2800–3150 cm−1 spectral range, which would result in omission of those particles from the final count. As such, MP/L concentrations reported in this study represent a conservative estimate. Intra-sample variability could not be assessed as only single samples were collected. Environmental factors such as temporal fluctuations in water flow, temperature, and biological activity could not be controlled or systematically assessed. These sources of uncertainty should be considered when interpreting both the concentration estimates and the broader implications of the findings.

    Data availability

    Data is provided within the manuscript or supplementary information files.
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    Download referencesAcknowledgementsThis work was funded by the Natural Sciences and Engineering Research Council of Canada (NSERC) (grant #576531-2022) and Health Canada (Environmental Health Research Contribution Program Agreement #2324-HQ-000112).Author informationAuthor notesThese authors contributed equally: Charles Balkenbusch, Judith Glienke.Authors and AffiliationsDepartment of Civil and Mineral Engineering, University of Toronto, Toronto, ON, CanadaCharles Balkenbusch, Judith Glienke, Yuhao Wu, Keenan Munno, Michael Jung, Husein Almuhtaram & Robert C. AndrewsAuthorsCharles BalkenbuschView author publicationsSearch author on:PubMed Google ScholarJudith GlienkeView author publicationsSearch author on:PubMed Google ScholarYuhao WuView author publicationsSearch author on:PubMed Google ScholarKeenan MunnoView author publicationsSearch author on:PubMed Google ScholarMichael JungView author publicationsSearch author on:PubMed Google ScholarHusein AlmuhtaramView author publicationsSearch author on:PubMed Google ScholarRobert C. AndrewsView author publicationsSearch author on:PubMed Google ScholarContributionsH.A. and R.C.A. conceived and designed the study. C.B., M.J., and H.A. developed the sample collection methodology. C.B., J.G., and M.J. collected samples and performed water quality analyses. C.B., J.G., and Y.W. performed Raman spectroscopy, with K.M. contributing to the Raman spectroscopy methodology. J.G. and K.M. performed statistical analyses. C.B., J.G., and H.A. wrote the first draft of the manuscript. J.G., K.M., H.A., and R.C.A. revised and edited the manuscript. R.C.A. provided supervision. All authors reviewed and approved the final manuscript.Corresponding authorCorrespondence to
    Husein Almuhtaram.Ethics declarations

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    Role of biofilm during groundwater biofiltration of manganese

    AbstractManganese (Mn) contamination in groundwater poses significant challenges for drinking water treatment. This study explores the mechanisms of Mn removal in a long-term oxygenated groundwater biofilter. The filter media coating primarily consists of abiotic disordered birnessite (δ-MnO2) with a microglobular structure and an average oxidation state of approximately 3.45. This material plays a key role in the effective adsorption and oxidation of Mn(II) dissolved in groundwater. The results indicate that Mn removal is predominantly abiotic, with biofilm activity contributing to less than 10% of Mn(II) oxidation. Biological colonization is minimal, as evidenced by the low microbial activity and protein-to-polysaccharide ratio. However, Mn-oxidizing and Mn-reducing bacteria were identified under aerobic conditions, suggesting that they play facultative or complementary roles in Mn cycling. The unexpected coexistence of the two types of bacteria highlights the need for further investigation into their role in δ-MnO2 transformation and regeneration. The study provides foundational insights into the dynamics of Mn(II) removal in biofilters and proposes an initial framework for understanding the Mn(II) biogeochemical cycle within such common engineered systems.

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    IntroductionManganese (Mn) is a contaminant commonly found in anaerobic groundwater, where it is mostly present as a reduced and dissolved cation (Mn(II)). Lowering Mn concentration in drinking water is a common drinking water treatment objective as elevated Mn concentrations may lead to aesthetic issues (colored water, metallic taste…)1 and neurotoxic effects on children2,3.Biological filtration is a commonly approved method for manganese removal, in which oxygenated groundwater is passed through a filter media such as sand to promote biofilm development. Over time, the filtration media becomes progressively coated with a layer of manganese oxides (MnOx) formed through the biogenic oxidation of Mn(II)4,5. This oxidation process is partly biotic, facilitated by various microorganisms6,7, and occurs predominantly through extracellular mechanisms8. These include direct oxidation catalyzed by proteins or polysaccharides, and indirect mechanisms involving changes in local environmental conditions5. The process is also partly abiotic, occurring through adsorption onto MnOx surfaces9.Bruins et al.4showed that MnOx (later identified as birnessite9,) is initially biogenic and gradually evolves toward abiotic precipitation/formation in less than 18 months. Mn removal efficiency by the media progresses over time, as demonstrated by the superior performance of a 15-year-old media compared to a 3-year-old media collected from biofilters10. Moreover, it was reported that Mn removal was reduced by 20–40% after eleven years of operation of a full-scale filter11.Although most often described as a biological process12, recent studies showed that Mn(II) removal is mainly abiotic13,14. In fact, Yang et al.14 tested multiples bacterial inhibition methods on a mature media to conclude that short-term Mn(II) removal was not impacted as long as the inactivation method did not alter the MnOx. The use of microbial inhibitors does not necessarily have an impact on the extracellular matrix already formed, which actively participates in the biological oxidation of Mn(II)5. Although biological and abiotic Mn(II) oxidation processes are known to be interrelated, maintaining MnOx reactivity over longer time span15, it remains important to accurately distinguish between biological and abiotic removal mechanisms within a biofilter in order to properly predict its performance.The biogeochemical cycle of Mn has been extensively studied in marine environments, with the identification of the coexistence of aerobic and anaerobic bacteria that can respectively oxidize Mn (MnOB)16 or reduce Mn (MnRB)17 or perform both roles depending on local redox conditions18. The co-existence of MnOB and MnRB has been observed in loose deposits from a chlorinated distribution network19 but this co-occurrence remains to be confirmed in aerated groundwater biofilters. Such a dynamic could influence treatment performance under shifts in water quality, as has been observed in riverbank filtration systems20. Understanding whether such microbial co-existence occur in biofilters is critical for predicting system stability and optimizing operational strategies under varying water quality conditions.The chemical Leucoberbelin Blue I (LBB) was shown to reduce MnOx to Mn(II)21, a characteristic which has been used with success to quantify the Mn oxidation activity of enzymes22. We hypothesize that biological Mn oxidation in a biofilter can be quantified after biofilm extraction by adding LBB in the presence or absence of sodium azide (NaN3) to discriminate the contributions of the MnOx and the cells present in the biofilm from the extracellular matrix. This is the first study to examine the Mn oxidizing capacity of a biofilm after its extraction from a biofilter media.This study aims to investigate the role of biofilm on Mn removal by characterizing the biofilter media surface, differentiating the mechanisms of Mn(II) oxidation within the biofilm, and confirming the presence of both MnOB and MnRB in the biofilm.Materials and methodsDescription of the water treatment plantThe pressure-driven biological filter under investigation has been in operation for over 20 years and feeds a small community in the southern part of Quebec (Canada), with a mean filtration velocity of 12.5 m/h. A description of the installation and the sampling procedure is presented in Text S1. The influent and effluent characteristics are presented in Table 1. As the influent contains no detectable iron, the sole objective of the treatment process is to remove approximately 0.4 mg/L of manganese, present primarily as Mn(II).Table 1 Characteristics of the biological filter influent and effluent.Full size tableBiofilm characterizationBiofilm can be defined as a microbial community embedded in an extracellular matrix composed of biological products such as extracellular polymeric substances, including MnOx which can be formed by biological oxidation of Mn(II). The biofilm, including its MnOx content, was extracted from the media sample by 3 cycles of 5 min sonication in phosphate buffer according to the method developed by Amini et al.23 which was further optimized for our biofilter media samples (Text S3).Biochemical characterization: Proteins were quantified at 560 nm with a micro-plate reader (Tristar2, Berthold Technologies) using the MicroBCA™ protein assay kit (Number 23235, Thermofisher®). Polysaccharides were quantified using the phenol method24,25,26. Metals (Mn, Fe, Ca and Mg) in the extracted biofilm solution were quantified by atomic absorption preceded by acid digestion with concentrated HCl27. Mn average oxidation state (AOS) in the extracted biofilm was determined using the LBB method21,28. More details about the methods used can be found in Text S4.Total/viable cells and diversity: Flow cytometry (BD Accuri™ C6 Plus, BDbiosciences) after LIVE/DEAD BacLight™ staining was performed in duplicate on the biofilm extraction water to assess bacterial viability29. Phenotypic diversity (alpha and beta) were determined from flow cytometry data using the Phenoflow method30.Bacterial identification: MnOB and MnRB were isolated by culture using Mn-oxidation31 and Mn-reducing32 agar media respectively with validation of the oxidation and reduction capacity performed according to the methodology of Cerrato et al.19. Pure strain of Pseudomonas putida (MnB1, ATCC) and Shewanella amazonensis (SB2B, ATCC) were used as positive control for the MnOB and MnRB medium, respectively.Extraction of genomic DNA was performed on nine MnOB, five MnRB isolates as well as 4 L of raw and filtered water and biofilm extract from the top of biofilter. Sequencing targeted the V4-V5 region of the 16S rRNA genes using the 515FB-926R primers33 on an Illumina MiSeq Sequencer (McGill Genome Center, Canada). Sequences were processed using the AmpliconTagger pipeline34. More details about the microbial quantification and identification are given in Text S5.Biofilm activityBiofilm activity was quantified using two analytical methods: ATP measurements to assess the general microbial activity and LBB oxidation kinetic to measure biological Mn oxidation activity.ATP analysis: Intracellular ATP quantification was performed on the biofilm extract using the protocol described in Text S2, while extracellular ATP was quantified on the filtrate to check the integrity of the biofilm extracted by sonication.Mn oxidizing activity: LBB reacts specifically with oxidized Mn to form Mn(II), and the number of electrons exchanged is directly proportional to the absorbance of the reactive solution at 630 nm21. The following methodology was developed to quantify Mn oxidizing activity of the biofilm extracts.Total biofilm oxidizing activity: The total Mn(II) oxidation activity of the biofilm (including MnOx) was determined in duplicate by monitoring for 8 h the oxidation of a Mn(II) solution by extracted biofilm (contact times: 0 min, 30 min, 1 h, 2 h, 4 h and 8 h). Absorbance measurements at 630 nm were performed on a solution containing 0.1 mL of the biofilm extract, 0.1 mL of a solution containing 1.5 mM of Mn(II) (MnSO4.H2O in phosphate buffer) and 0.1 mL of 0.04% (w/v) LBB spiked at each contact times (in different test wells). Inhibited microbial oxidizing activity: Samples were spiked with NaN3 (Cfinal = 0.015M) to reduce the activity of the respiratory chain. Then the LBB solution was spiked in the test wells prior to the 8-h monitoring of absorbances as described above. As LBB rapidly dissolves MnOx, Mn oxidation was assumed to only result from other components of the extracellular matrix. The complete reduction of MnOx by LBB was validated by the absence of changes in the absorbance of controls experiments without MnSO4.For both types of assays, LBB was measured using the analytical method described in Text S4. A summary of the differences between the three types of Mn oxidation activities is presented in Table S1. Influence of the addition of LBB on the biofilm activity is discussed in Text S4.Biogenic manganese removal modelingKinetic data modeling: The LBB analysis quantifies electrons exchanged to Mn(II). However, to determine the rate of Mn(II) oxidation by the biofilm, it is necessary to select the number of electrons exchanged during Mn(II) oxidation by the biofilm. In this study, we assumed an average oxidation state (AOS) for biogenic MnOx of 3.5 (3.5 implies the exchange of 1.5 electron for each Mn(II) being oxidized) (Eq. 1).$${Mn}^{+2}+{O}_{2}underset{biofilm}{to }{{Mn}^{+3.5}O}_{2}$$
    (1)
    Mn(II) oxidation during LBB assays were described by a pseudo-first-order kinetic model according to Eq. 2. The model parameters f and k were used to compare the samples with total biofilm activity and the residual biofilm activity after inhibition of the respiratory chain (inhibited microbial activity) which corresponds to the residual activity of the extracellular matrix. The parameters k and f respectively describe the oxidation kinetic (in h−1) and the maximum MnOx oxidative capacity (mg Mn oxidized/cm3 of media).$$left[{{Mn}^{+3.5}O}_{2}right]=ftimes left(1-{e}^{-kt}right)$$
    (2)
    Maximal Mn oxidized by the biofilm in a full-scale biofilter: The maximum amount of Mn(II) that could theoretically be oxidized by the biofilm in our industrial biofilter was estimated using the factor f, determined above for different media thickness, which corresponds to the amount of Mn oxidizable by the biofilm per volume of media. Given that this factor was determined at three media depths, the maximum of Mn(II) oxidized in the whole filter was calculated using a weighted average (Eq. 3). This value was compared to the total Mn removed by the biofilter assuming continuous biofilter operation at 12.5 m/h with Mn(II) concentration of 0.37 mg/L in the influent.$$Max Mnleft(IIright) removed by biofilm (frac{g Mn}{kg media})=sum_{for each layer}f*frac{Volume}{media mass}$$
    (3)
    In practice, the removal will be limited by the kinetic due to the short contact time available in a biofilter. Assuming a packed bed with negligible dispersion and advection, Eq. (4) was used to estimate the removal by the biofilm while considering kinetic limitation:$$frac{left[Mnleft(IIright)right]}{{left[Mnleft(IIright)right]}_{0}}={e}^{-kt}$$
    (4)
    where k is the pseudo-first order kinetic parameter determined earlier (h−1), and t (in h) is the estimated real contact time assuming a filter bed porosity of 45%.Characterization of filter mediaThe metal content (Ca, Fe, Mg and Mn) of the media was determined by atomic absorption spectroscopy using a PinAAcle 900 F spectrometer (PerkinElmer) which was preceded by acid digestion with HCl27. The abiotic vs. biotic origin of the media was determined by electron paramagnetic resonance (EPR) at a frequency of 9.6 GHz35 according to the reference values proposed by Kim et al.36.The surface of the coating was studied by environmental scanning electron microscopy (ESEM) (Quattro, Thermofisher) coupled to an energy dispersive x-ray spectroscopy (EDS) detector (Ultim-Max, Oxford Instruments). The depth of the media was studied with a scanning electron microscopy (SEM) (FEG-SEM JSM-7600F™, Jeol®) coupled with an EDS detector (X-max™, Oxfort Instruments) using cut, coated and polished media grains.The mineralogical nature of MnOx was identified by X-ray diffraction at the European Synchrotron Radiation Facility on beamline ID2237. Measurements were performed using standard borosilicate glass capillaries in Debye–Scherrer geometry, Si(111) monochromated beam (E = 35 keV), and a Perkin Elmer XRD1611 detector. The detector geometry was calibrated with the LaB6 standard. Mn K-edge X-ray absorption spectra were collected at the Canadian Light Source (CLS) synchrotron on beamline IDEAS (08B2-1) as described in Text S6. Data preprocessing and analysis were performed with Athena38. Mn AOS in the solids were determined using the Combo Method based on Mn(II), Mn(III) and Mn(IV) proportions39 and a Mn reference database40.Statistical analysisComparisons between results at different depths were performed using the Kruskal–Wallis test using R41 a p-value < 0.05 was considered as significant.ResultsMedia characterization and confirmation of abiotic MnOx with biogenic-like growth structuresFigure 1 illustrates the ESEM imaging of the media structure. The media exhibits remarkable homogeneity, characterized by a repetitive globular microstructure resembling coral or sponge-like formations (with Mn and O as major elements), which are typical of abiotic MnOx4,42. The presence of overlapping globules, loosely connected to one another, suggests that media growth occurs at specific localized sites. The abiotic origin of the samples from various media depths was confirmed through EPR analysis (Table S2), which showed a minimum linewidth (ΔH of 1870 Gauss (187mT > 1200 Gauss (120mT36;). In Fig. 1A and 1B, small particles (indicated by arrows) can be observed directly attached to the globular abiotic MnOx. Under higher magnification (Fig. 1C and 1D), a fluffy structure typical of a layered MnOx becomes visible which as been reported as biotic4,42. Although abiotic MnOx dominates the structure, some oxides at the growth centers appear to have a biological origin based on the fluffy structure.Fig. 1Environmental scanning electron microscopy of media surface at different magnifications (500X for A, 2,500X for B and 10,000X for C and D). Arrows point out suspected biological structures.Full size imageThe chemical composition of the coating, considering Mn, Fe, Ca, and Mg (Table S2), is over 85% Mn and remains consistent throughout the entire depth of the filter (p > 0.05). A cross-section of a media grain (Fig. 2A) reveals distinct structural features between the surface and the deeper layers of the coating, as well as the presence of interstices, which are further confirmed by SEM imaging (Fig. 2B). The chemical composition of the surface coating analyzed using EDS (with a penetration depth of 0.8 µm in MnO₂ at 10 keV43) shows a decrease in the relative Mn content as a function of depth in the biofilter (Fig. S6). This indicates notable changes in the chemical composition between the surface of the coating and the bulk of the filter media. Mn(II) adsorption is more pronounced at the top of the biofilter, as indicated by the lower O/Mn ratios of the bulk coating in the upper section compared to the lower section (Fig. S7).Fig. 2(A) Microscopic observation of a cross-section of a filter media grain, with the support media (SiO2) in light grey, the MnOx coating in black, and the polymer used for the cross-section in white. (B) SEM observation of the filter media cross-section, in black the polymer used to prepare the media grain cross-section, the light grey area corresponds to the MnOx covering of the media and the dark grey area is the SiO2 core of the filter media.Full size imageBecause the overall chemical composition of the media did not vary much with depth or after backwashing, only two samples at the top of the biofilter, at the beginning (S1) and at the end (S2) of the biofiltration cycle, were analyzed for Mn AOS using XANES spectroscopy and characterized structurally using XRD and EXAFS spectroscopy. The AOS of Mn is 3.45 ± 0.04 in S1 and 3.56 ± 0.04 in S2 with Mn(III) amounts to 30–35% and Mn(II) amounts to 7–10% (Table S2 and Fig. S8). Mn AOS values are on the low end of those reported previously for a biofiltration pilot system (3.5 and 3.99;). The XRD patterns of both samples are characteristic of turbostratic birnessite (δ-MnO₂) mixed with quartz (Fig. S9), consistent with previous studies4,9,44,45. The d-spacings of the (100) and (110) reflections are 2.456 Å and 1.419 Å. Their ratio is 1.731, which indicates that the symmetry of the MnOx layer is hexagonal. For both samples, the EXAFS spectra are similar, and their shape is also characteristic of hexagonal phyllomanganate (Fig. S10). Of significance is the shape of the so-called “indicator region” in the 7.5–8.5 Å−1 k interval46,47. When the layer symmetry is hexagonal, the electronic wave is a single antinode and a double antinode when it is orthogonal (Fig. S10.A). Departure of the hexagonal symmetry occurs when the Mn(III) and Mn(IV) cations in the MnO2 layer are ordered in rows (Fig. S11). Sample spectra show that the antinode is asymmetric on the left side, which seems to support a small lowering of the layer symmetry. This observation is consistent with the high amount of Mn(III) and Mn(II) cations in the MnOx structure, part of them being probably within the layer. The distortion of the layer structure caused by the Mn(III)/Mn(IV) ordering should be local, because EXAFS is a short-range probe, in contrast to XRD. The emergence of a Mn(III) ordering, together with the low AOS, suggest that the solid gradually transforms into a thermodynamically more stable arrangement over time, such as feitknechtite48. This indicates that the coating undergoes very long-term changes.Microbial and biochemical characteristics of the extracted biofilmTable 2 presents the biochemical composition of the extracted biofilm at the beginning and at the end of the biofiltration cycle. Unlike the typical biofilm stratification observed in activated carbon biofilters49, the depth-dependent analysis (Table S3) in this study reveals a weak influence of depth on biofilm composition, with only a significant increase in polysaccharide concentration observed with depth (p < 0.05). The biofilm ATP concentration (from 60 to 91 ng/cm3) was slightly above values reported by Keithley et al.50 for samples collected at the top of two Mn biofilters fed with groundwater (~ 35 ng/cm3) but was significantly lower than the ATP concentration reported by McCormick et al.51 (~ 250 ng/cm3, converted considering a sand density of 2.7) for a Mn biofilter also fed with groundwater. Protein content and the protein-to-polysaccharide ratio in this study were notably low compared to those reported in the literature50,51. These low protein levels suggest a cohesive biofilm structure that promotes metal retention52, as evidenced by the significant Mn content (22–29 mg/cm3) with an AOS of 3.36 and 3.51 confirming the presence of MnOx in the biofilm matrix.Table 2 Biochemical composition of the biofilm in the top of the biofilter at the beginning and at the end of the filtration cycle.Full size tableDuring the filtration cycle, the overall biological activity (lower concentration of polysaccharides and intracellular ATP) of the biofilm decreases significantly (p < 0.05), indicating that backwashing restores the biological activity lost during the filtration cycle.Fig. S12 shows a comparison of the microbiological diversity present at the top of the biofilter, in the influent and in the effluent. Most bacterial class are less abundant in the top of the filter than in the influent or effluent, except in particular for the Gammaproteobacteria and Alphaproteobacteria which have already been identified as the groups most commonly found in groundwater-fed biofilters50. This observation suggests that only a fraction of the influent bacteria can proliferate within the biofilm of the biofilter. The most abundant bacterial genus in the media also belong to the Gammaproteobacteria and Alphaproteobacteria classes, and correspond to the X35OR, Sphingopyxis and Coxilla genus.Multiple MnOB were isolated by culture, as expected considering the aerobic conditions prevailing in the bioreactor, however unexpectedly multiple MnRB colonies were also isolated from most media samples by aerobic culturing. While several bacterial genera were isolated from MnOB cultures (Table S4), only the Brevundimonas genus (found in both MnOB and MnRB cultures, Table S5), was identified in MnRB cultures, along with the Bacillus genus, which includes multiple strains previously reported to oxidize manganese19. These results highlight the presence of facultative Mn reducers in the biofilter, which could cause Mn release under certain conditions, such as a temporary filter shutdown.Limited Mn oxidation by the biofilm extractFigure 3 presents the variation of the extracted biofilm oxidative activity parameters (f and k determined using Eq. 2) in function of filter depth. The total biofilm oxidative activity kinetic constant (k) for each depth was used to estimate the biological contribution to Mn(II) removal in the industrial biofilter using Eq. (4). From this calculation, the maximum amount of Mn(II) removed biologically, based ona constant water velocity of 12.5 m/h, is estimated as ≈ 4% of the initial Mn(II) concentration for the biofilter investigated. According to the kinetic study, a minimum contact time of 6 h would be required to remove 99% of the Mn concentration by direct oxidation by the biofilm alone. Mn(II) removal is therefore predominantly abiotic, as confirmed by the study of the elemental composition of the media surface at the top of the filter demonstrating strong adsorption of Mn(II) (Fig. S7). However, once Mn(II) is adsorbed onto the media, it remains available for subsequent oxidation catalyzed by the biofilm or the surface media, independent of water contact time.Fig. 3Evolution of the model parameters (Fig. 3A: f, Fig. 3B: k) for Mn oxidation ((□) Total biofilm, (Δ) microbial with inhibition of respiratory chain) in function of depth, at the beginning of the filtration cycle. Fig. S13 presents the same results at the end of the filtration cycle. Fig. S14 presents the raw data of a kinetic assay.Full size imageThe maximum Mn(II) oxidation capacity of the biofilm (factor f) was estimated to be approximately 14.5 g Mn/kg of media in the industrial biofilter at the beginning of the cycle. After two weeks of operation, corresponding to the removal of approximately 0.38g Mn/kg of media, a 36% reduction in the maximum capacity (equivalent to 5.2 g Mn/kg) was observed, with the middle section of the biofilter experiencing a more pronounced loss of 47%. The biofilm’s maximum Mn(II) oxidation capacity is sufficient to handle the influent concentration of 0.37 mg Mn(II)/L entering the biofilter. However, under actual operating conditions, the short contact time and competition with other compounds that can also be oxidized by the biofilm prevent the full utilization of this capacity.The highest Mn oxidation capacity (f = 48 mg/cm3) was observed in the total biofilm extract (MnOx + bacteria), suggesting that the presence of MnOx and bacteria within the biofilm significantly enhances Mn(II) uptake (p < 0.05). Yet, an average of 36 ± 12% of Mn oxidation capacity persisted after the addition of NaN3 and LBB, which corresponds to the oxidation capacity of the extracellular matrix without MnOx. Extracellular polymeric compounds play a non negligible role in the MnOx oxidation by the biofilm especially in the upper layer of the biofilter. Moreover, the oxidation kinetic constant (k) was significantly higher (p < 0.05) in the biofilm sample without MnOx compared to the other pathway. This may indicate that the presence of MnOx introduces an additional adsorption step, which increases the overall Mn(II) removal capacity but reduces the apparent oxidation rate.DiscussionThe kinetic study of biological Mn oxidation predicted that up to 4% of Mn can be directly oxidized biotically when the biofilter operated at a flow rate of 12.5 m/h (according to Eq. (4)). Our approach based on measuring the oxidation rate of extracted biofilm to determine the maximum biological removal yields an estimate which is consistent with those derived using biofilm inactivation techniques without extraction13,14. However, the study of biofilm oxidation capacity shows that the contribution of the extracellular matrix is not negligible compared with that of the biofilm and the extracellular activity is not inhibited during biofilm inactivation by NaN3. The biofilm present in this biofilter exhibits limited activity but contains a high concentration of polysaccharides, which are known to adsorb Mn(II) and serve as nucleation centers for biological MnOx formation53. The presence of such nucleation zones was confirmed by ESEM analysis of the media samples.The investigation into the biological mechanisms responsible for Mn oxidation showed that the presence of MnOx in the biofilm tends to reduce the biological oxidation rate. MnOx present in the biofilm can be seen as a competitor with extracellular enzymes for reacting with Mn(II). Nevertheless, the total biofilm first-order Mn(II) oxidation rate was 3.4 times higher than the rate we derived for the abiotic oxidation of Mn(II) using biogenic MnOx54. This higher reactivity of biotic MnOx is aligned with observations from marine systems15.In the investigated biofilter, the MnOx coating is predominantly abiotic. Although abiotic Mn(II) oxidation is slower than biological oxidation, the prevalence of Mn(III) and loss of hexagonal symmetry in the solid matrix confirm its role in regenerating autocatalytic sites and indicates that MnOx coating continues to evolve over time. Biological regeneration, though minor, helps maintain more reactive higher valency MnOx. Adsorption is the primary Mn(II) removal mechanism, but sufficient dissolved oxygen (DO) is critical for regenerating adsorption sites and ensuring long-term biofilter performance. Future studies should explore the influence of DO on birnessite regeneration.Biological MnOx oxidation is known to be mostly an extracellular process8. In our biofilter, 36% of the capacity is directly related to polysaccharides and proteins which remain in the extracellular matrix and play a major role in the biological Mn oxidation.Although Mn reducing activity has not been proven in this biofilter, the presence of bacteria with the capacity to reduce Mn (MnRB) may lead to the release of Mn(II) in the biofilter effluent in the event of a prolong change in redox conditions.The initially biological MnOx coating becomes abiotic over time4, and evolves towards a thermodynamically more stable form, which may modify these properties and therefore influence Mn retention over the very long-term. This change, reported for the first time in this study, requires further investigation to understand its implications for the biogeochemical cycle of Mn.Figure 4 illustrates the proposed biogeochemical cycle of Mn in an aerated biofilter, where Mn uptake during biofiltration is predominantly abiotic. Mn(II) in the liquid phase is primarily adsorbed onto the media surface (step 1 A), with a smaller fraction retained on the biofilm’s extracellular matrix (step 1B). Mn(II) is then oxidized either autocatalytically by MnOx coatings (step 2 A) or bio-catalytically (step 2B, minor contribution). Abiotic Mn(II) oxidation produces solid Mn(III, IV), which is incorporated into the media coating structure5. Although the role of oxygen in this step is not fully studied, it likely contributes to regenerating MnOx by accepting electrons during the oxidation of inorganic contaminants55,56,57, consistent with the observed decrease in AOS. Newly formed MnOx sites can then catalyze further Mn(II) adsorption (step 3 A).Fig. 4A proposed biogeochemical Mn cycle in a manganese biofilter. The dotted arrows indicate progressive steps. Mn(II) was represented in blue in bulk water, green when sorbed and in brown when oxidized. Created using Biorender.com.Full size imageBiocatalytic Mn(II) oxidation generates Mn(IV) solids, with dissolved Mn(III) as an intermediate trapped in the biofilm matrix (steps 2B and 3B)58,59. Polysaccharides in the biofilm serve as nucleation centers for MnOx, facilitating media growth60,61,62. Biological Mn(II) oxidation is often mediated by extracellular multicopper oxidases (MCOs)5,58, although reactive oxygen species may also play a role15,63,64.ConclusionThis study provides comprehensive insights into the mechanisms responsible for long-term Mn removal in groundwater biofilter. The coating on the filter media was identified as predominantly abiotic disordered birnessite (δ-MnO2) with a globular microstructure and an AOS of approximately 3.5. This microglobular structure reflects the progressive growth of abiotic Mn oxides, which are primarily responsible for Mn(II) adsorption and oxidation. The MnOx coating slowly evolves towards a thermodynamically more stable state.Manganese removal in the investigated biofilter was mainly driven by abiotic processes, with the biofilm contributing less than 10% to the overall Mn(II) oxidation. The biofilter exhibited low biofilm colonization, with limited microbial activity, as evidenced by the low ATP concentrations and low protein-to-polysaccharide ratios. Despite this, the biofilm indirectly supports the regeneration and maintenance of MnOx deposits, facilitating sustained Mn removal performance.Both MnOB and MnRB were detected under aerobic conditions, suggesting potential facultative behavior or complementary roles in Mn cycling. The presence of facultative MnRB in the biofilter could lead to Mn release in the event of a change in biofilter operating conditions (e.g. long stagnation time, drop in dissolved oxygen, etc.).Looking ahead, further research is needed to clarify the role of biofilm in the long-term regeneration of filter media, particularly in relation to dissolved oxygen levels. A key question emerges: are microbial communities essential for the sustained functionality and self-regeneration of a mature biofilter, or could abiotic processes alone be sufficient once sufficient coating has been produced? Addressing this question could guide strategies to optimize biofilter performance and ensure its long-term efficacy in Mn removal.This study emphasizes the predominance of abiotic mechanisms in Mn biofiltration while acknowledging the potentially supportive role of microbial communities in maintaining the long-term reactivity of filter media.

    Data availability

    The data presented in this study are available from the corresponding author upon reasonable request. The raw sequence reads for 16SRNA sequencing are available in the European Nucleotide Archive (ENA) database under project number PRJEB95825 (https:/www.ebi.ac.uk/ena/browser/view/PRJEB95825).
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    Download referencesAcknowledgementsThis study has been supported by the Natural Sciences and Engineering Research Council of Canada (NSERC) through its Discovery Grant program (RGPIN2020-0649). Part of the research described in this paper was performed at the Canadian Light Source (CLS), a national research facility of the University of Saskatchewan, which is supported by the Canada Foundation for Innovation (CFI), the Natural Sciences and Engineering Research Council (NSERC), the Canadian Institutes of Health Research (CIHR), the Government of Saskatchewan, and the University of Saskatchewan. The authors thank the staff of the municipality who let us access and study the installation described in this article. The authors thank David Muir for his welcome and grateful help to perform XAFS analysis at the IDEAS beamline (CLS) and, Jacinthe Mailly, Julie Philibert, Mélanie Rivard, Tetiana Elyart and Yves Fontaine for their help to perform the coring and the experimental analysis.FundingThis study has been supported by the Natural Sciences and Engineering Research Council of Canada (NSERC) through its Discovery Grant program (RGPIN2020-0649). Part of the research described in this paper was performed at the Canadian Light Source (CLS).Author informationAuthors and AffiliationsDepartment of Civil, Geological and Mining Engineering, Polytechnique Montreal, 2900 Edouard Montpetit, Montreal, QC, H3T 1J4, CanadaJérôme Ducret & Benoit BarbeauEuropean Synchrotron Radiation Facility (ESRF), 38043, Grenoble, FranceAlain Manceau & Catherine DejoieLaboratoire de Chimie, ENS de Lyon, CNRS, 69342, Lyon, FranceAlain ManceauDepartment of Engineering Physics, Polytechnique Montreal, 2900 Edouard Montpetit, Montreal, QC, H3T 1J4, CanadaChristian Lacroix & David MénardAuthorsJérôme DucretView author publicationsSearch author on:PubMed Google ScholarAlain ManceauView author publicationsSearch author on:PubMed Google ScholarChristian LacroixView author publicationsSearch author on:PubMed Google ScholarDavid MénardView author publicationsSearch author on:PubMed Google ScholarCatherine DejoieView author publicationsSearch author on:PubMed Google ScholarBenoit BarbeauView author publicationsSearch author on:PubMed Google ScholarContributionsJD and BB performed the conceptualization, methodology, the collection and the formal analysis of the data set excluding XRD, XANES, EXAFS and EPR analysis, they wrote the main manuscript. AM performed the XANES, EXAFS and XRD data analysis. CL and DM performed the EPR analysis and data collection. CD performed the XRD data collection. All authors reviewed the final manuscript.Corresponding authorCorrespondence to
    Jérôme Ducret.Ethics declarations

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    Reprints and permissionsAbout this articleCite this articleDucret, J., Manceau, A., Lacroix, C. et al. Role of biofilm during groundwater biofiltration of manganese.
    Sci Rep 15, 41330 (2025). https://doi.org/10.1038/s41598-025-25228-5Download citationReceived: 07 August 2025Accepted: 18 October 2025Published: 21 November 2025Version of record: 21 November 2025DOI: https://doi.org/10.1038/s41598-025-25228-5Share this articleAnyone you share the following link with will be able to read this content:Get shareable linkSorry, a shareable link is not currently available for this article.Copy shareable link to clipboard
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    Elucidating the loose tie between precipitation and streamflow sensitivities to warming across the contiguous United States

    AbstractThe relation of regional precipitation and streamflow sensitivities to warming is complicated by runoff often being the small residual in the balance between precipitation and evapotranspiration. In addition, the inhomogeneous nature of land-hydrological properties and hydrological interventions by humans, like irrigation and dams, have made streamflow sensitivities challenging to constrain with both climate models and observations. Here, we elucidate the hydrological processes driving the regionally varying mean and extreme streamflow sensitivities to warming across the contiguous United States in a counter-factual warming scenario using GFDL’s moderately high-resolution global climate model. We show that over the West Coast and eastern US, atmospheric rivers are the dominant driver of high-flows in the present climate and of more frequent high-flows in a warmer climate. Over the mountainous western US, streamflows dwindle despite increases in precipitation and antecedent soil moisture, as loss of snow fuels evapotranspiration at double the rate of precipitation changes.

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    IntroductionUnderstanding how regional precipitation and river floods respond to increasing temperatures is a crucial question for climate science, as it has severe implications for the mitigation of natural hazards that cost lives and property1. In the United States, the contribution of historical precipitation changes to flood damages has been estimated to be an accumulative $73 billion2, highlighting the need for a better understanding of the relation of regional precipitation and flood changes under warming. At global scales strong constraints on precipitation changes with warming exist with an expected mean increase of 2–3% K−1 3,4,5,6, large-scale patterns following the paradigm of wet gets wetter, dry gets drier3,7,8,9 and an overall shift towards more extreme and less moderate precipitation events10,11,12,13,14,15,16,17; a change in the character of precipitation18. However, at regional scales vast uncertainties remain as precipitation changes over land are more complex than wet gets wetter, dry gets drier19 and a computational need remains for compromising between spatial resolution of climate models, long integration times and ensemble or domain sizes16,20,21,22,23, as well as limitations in process understanding6,24,25,26.While these constraints also limit predictability of changes in river floods with warming, precipitation is only one of a multitude of land-hydrological processes that determine flooding and its change with warming27,28,29,30. The amount of precipitation necessary to produce runoff is strongly determined by antecedent soil conditions, which are controlled by processes such as previous precipitation, snow-melt and evapotranspiration (ET)31,32,33,34 that in climate models often require strong assumptions about the sub-grid scale. In addition, observational constraints on trends in river floods remain limited due to weak signal-to-noise ratios and human interventions35,36. Hence, improving global coupled atmosphere-land models to higher resolutions offers an exciting opportunity to enhance our process understanding of precipitation and flood changes with warming.Recent advances to moderately high resolutions of 50 km and higher driven through the CMIP6-HighResMIP initiative37 have shown promising improvements in representing precipitation and its extremes38,39,40,41,42,43,44, particularly in the mid-latitudes where mesoscale frontal systems often dominate the spatial scales of precipitation45,46. A strong contributor to mid-latitude precipitation and floods are atmospheric rivers (ARs)16,32,47,48,49,50, which are characterized as elongated narrow bands of intense poleward moisture transport and typically associated with extra-tropical cyclones47,51,52. Previous studies point to a substantial improvement in the representation of AR precipitation in moderately high-resolution models and highlight AR precipitation changes with warming that account for a significant fraction of the total mid-latitude precipitation response, especially over the United States45,46,49,53,54,55,56. Despite this, initial assessments of moderately high-resolution models for hydrological applications are limited. Recent first results show promise in capturing peak annual runoff over the central US, with needs for downscaling and bias correction remaining to accurately assess local scales57.Here, we further explore the new opportunities provided by moderately high-resolution to assess changes in river floods and their drivers with warming, distinguishing AR and non-AR precipitation, as well as snow melt, across the contiguous United States (CONUS). We utilize the global coupled atmosphere-land model AM4.0/LM4.058,59 with a resolution of 50 km46 developed at the Geophysical Fluid Dynamics Laboratory (GFDL), which has shown promising capabilities in capturing mean and extreme precipitation45,46,58. LM4.0 discharges surface and sub-surface runoff across a grid-scale river network (Supplementary Fig. S1) as fluxes of liquid water, ice and heat to connected grid-cells60,61. This approach benefits from the grid-scale coupling of hydrological processes across river, land and atmospheric domains, allowing for a definitive attribution of various hydrological drivers of high river streamflows indicative of flooding and hence a process-based understanding of their sensitivity to warming.Our experimental setup (detailed in the section “Model setup and simulations”) is designed to isolate the best constrained components of precipitation changes with warming, providing a sound basis for assessing the relation of changes in precipitation and streamflows. First, we isolate the thermodynamic precipitation response to warming, which excludes any changes in atmospheric winds and denotes the most significant and well-understood part of how precipitation extremes12,24,62 and AR precipitation49,63,64,65 change with warming. We achieve this by tightly nudging atmospheric winds toward NCEP reanalysis data66 for both a control and a counter-factual warming simulation that we run over the historical period of 1951 to 2020. The control simulation (CTRL) follows the CMIP6 HighResMIP protocols37 and is forced by observed daily sea surface temperatures (SSTs). The counter-factual warming simulation only differs in SST forcing by a homogeneous increase of SSTs across the global ocean by 2 Kelvin (P2K). Besides isolating the most significant and well-constrained parts of the total precipitation response to warming, our nudging approach also minimizes uncertainty arising from natural climate variability as well as dynamically driven model biases (see Supplementary Fig. S2), enhancing applicability to hydrological applications. In addition, we obtain a day-by-day comparability between the CTRL and P2K simulations as well as with observations (Supplementary Figs. S3 and S4), which we exploit in the section “Relative importance of changes in antecedent upstream soil moisture and water inputs”.To elucidate the relationship of precipitation and streamflow changes with warming across CONUS, we first quantify their sensitivities to warming on the mean and in the extremes in the section “Precipitation and streamflow changes with warming”. We then identify high streamflows and investigate changes in their frequency with warming across CONUS, which we characterize by examining three inter-related conditions for high-flows. Firstly, we attribute high-flows to different upstream water inputs and their changes with warming in the section “Attribution to upstream precipitation and snow melt”, distinguishing AR and non-AR precipitation as well as snow-melt (as described in the section “Attribution to upstream precipitation and snow melt”). Secondly, we examine changes in water and energy constraints on losses through ET and runoff in the section “ Attribution to seasonal constraints on evapotranspiration”, highlighting seasonal constraints on the high-flow response. Thirdly, we discuss how these constraints are imprinted on antecedent soil moisture conditions prior to high-flows and to what degree these explain changes in high-flows compared to changes in upstream water inputs in the section “Relative importance of changes in antecedent upstream soil moisture and water inputs”. We draw final conclusions in the section “Discussion”.ResultsPrecipitation and streamflow change with warmingFigure 1 summarizes the thermodynamically driven changes in precipitation and streamflows across CONUS with warming in our counter-factual warming experiment. The thermodynamic precipitation response manifests as an overall increase across most of CONUS with an average of 3.6% K−1 (with respect to the global mean warming signal in 2 m temperature), except in the dry southern states Arizona and New Mexico (Fig. 1a). To give some reference to this number, observed scaling of annual mean precipitation over CONUS between the years 1901 and 2015 is estimated at 4%67, which given an increase in global mean 2 m temperature in that time of about 1.1 K68 translates to a scaling of exactly 3.6% K−1. Given the idealized assumptions of our experiment about constant atmospheric winds and the ignored direct effects of greenhouse gases20,69, we have to acknowledge that we obtain the same number for different reasons. However, comparability to the observed mean precipitation response increases the applicability of our results.Fig. 1: Regional contrasts in precipitation and streamflow sensitivities to warming.Changes with warming in a mean precipitation, b extreme precipitation, c mean AR precipitation, d mean streamflow and e extreme streamflow across CONUS based on counter-factual warming experiment (see the section “Model setup and simulations”). “Extreme” refers to the 99.7th percentile of daily values, resembling events with an annual recurrence rate, also indicated by black vertical line in (f). Hatched regions denote insignificant changes of the mean as determined by z-test with critical value z = 1. f Shows relative change with warming in precipitation, AR precipitation and streamflow deduced for each grid-point as function of percentile and then averaged across CONUS (black) and for streamflow also across the west coast (green), the mountainous west (red) and the eastern US (blue) defined by zonal bounds at 120°W and 100°W, respectively, indicated by gray dashed lines in (e). Shading denotes standard error across respective region. Red solid line indicates Clausius–Clapeyron (CC) scaling of 7% K−1.Full size imageTo consider changes in extreme precipitation, we define extreme precipitation as the 99.7th percentile of daily precipitation which represents an average recurrence rate of once per year as commonly used in studies of hydrological extremes70,71,72,73. Changes in extreme precipitation (Fig. 1b) are positive across CONUS and at 5.4% K−1 generally larger than mean changes, with many local exceedances of the Clausius-Clapeyron (CC) scaling at 7% K−1 18,74,75, highlighting the expected shift towards more intense precipitation10,11,12,13,14,15,16,17. Mean AR precipitation changes (Fig. 1c, see the section “Atmospheric river detection and evaluation” on AR detection) fall between mean and extreme precipitation changes at an average of 4.2% K−1. We want to emphasize that these scalings are not contradicting previously reported super-CC scaling of AR precipitation76,77, which stems from a focus on sub-daily timescales while here we focus on daily timescales. We do, however, find a strong regional pattern in extreme AR-precipitation scaling with warming, with some local exceedances of CC-scaling across the mountainous west, somewhat resembling the lee-side of the Sierra Nevada, the Cascades and the Rocky Mountains. These results indicate a weakening of the orographic rain shadow with warming, which has previously been predicted from theory78 and high-resolution regional modeling studies79,80,81 and is therefore encouraging to identify in our moderately high-resolution simulations.The warming response of mean and extreme river streamflows (Fig. 1d and e) strongly differs in its pattern and magnitude from the precipitation response, reflective of the streamflow response being controlled by a multitude of hydrological processes such as antecedent soil conditions and ET31,34,54,82,83,84,85,86. Most prominently, mean and extreme streamflows decline over the mountainous western US (regions indicated by gray dashed lines in Fig. 1e and colored text below) in the presence of significant precipitation increases, particularly from ARs. Streamflow declines under warming have previously been noted and linked to the dominance of spring-time snow melt in driving hydrographs29,30,72,87,88,89. Other, more regional studies have identified pronounced increases in evaporative demand90 and ET91 across the southwestern US as drivers of drought and dwindling streamflows. Outside the mountainous west, streamflows increase with precipitation; however, the increase is significantly stronger over the east than over the west coast (Fig. 1e, note that we consider large parts of California towards the mountainous west, as we orient our regional definition based on the streamflow response).We summarize the regional discrepancies of extreme precipitation and streamflow changes with warming across sub-regions of CONUS in Fig. 1f through mean relative changes in percentiles (following previous studies20,34,92). The graphs can be interpreted as the regional mean change of daily precipitation or streamflow at a given recurrence rate described through the percentile on the x-axis (e.g. annual recurrence resembled through the 99.7th percentile of daily values). We find that precipitation changes show an expected increased sensitivity to warming at higher percentiles6,12,20, with AR precipitation being more sensitive than total precipitation45,55,56,93, approaching CC-scaling for the strongest extremes. Streamflows, however, show a much different behavior in their scaling across percentiles, representative of their loose tie to precipitation. On the one hand, we find an exceedance of precipitation changes by streamflow changes over the eastern US across all percentiles, which exceeds CC-scaling at percentiles >98. Over the mountainous west, on the other hand, streamflows dwindle across the distribution and, even at the highest percentiles, where precipitation increases the most. Streamflow changes over the West Coast lie between the other two regions, with generally positive changes but at a lower magnitude than CONUS-averaged precipitation changes.Changes in the surface water balanceWe can understand the regional variation in streamflow sensitivities to warming by considering the regionally integrated surface water balance and its changes with warming. The local surface water balance can be expressed as$$P=ET+R+Delta S$$
    (1)
    At any location, the local input of water through precipitation P is balanced by losses through ET, runoff R and a change in water storage on land ΔS. The local river streamflow is determined by the streamflow from upstream and the local runoff; in our model, the river does not feed water back to its surrounding soil59,61. Figure 2a depicts the regionally and annually averaged terms of the surface water balance in the CTRL simulation (circles) normalized by the regional mean P. For reference, Fig. 2a also shows observationally deduced estimates of the regional surface water balance (crosses) obtained from a 1 km resolution Thornthwaite-style water balance model94 that is forced by daily observations of precipitation and temperature95,96,97 (detailed in the section “Evaluation of the surface water balance”).Fig. 2: Regional differences in surface water balance.a Shows regional annual average (1980–2020) terms of the surface water balance (Eq. (1)) in the CTRL simulation (circles) and observationally deduced based on a surface water balance model95 (crosses), normalized by the regional mean precipitation P to make different regions comparable. b Shows the respective change with warming in each of the terms, normalized by the global mean warming response. Note that changes in b show absolute changes of the budget terms in units of % K−1 due to normalization by P. Errorbars denote spatial standard deviation within the respective regions.Full size imageWe find that the partitioning of ET and R across the three regions is vastly different, with R/P denoting only 18% over the mountainous West, 28% over the eastern US and 51% over the west coast. Over the mountainous west and eastern US, these numbers are in remarkably good agreement with the observationally deduced estimates, while over the west coast, R is underestimated by our model, and ET is overestimated. We attribute this bias to overestimated ET/P over major mountain ranges, such as the Sierra Nevada and the Cascades, which are not sufficiently resolved in our 50 km simulations (Supplementary Fig. S5). Underestimated orographic lift yields underestimated P in mountainous terrain, where ET is generally less efficient due to enhanced partitioning to snow and lower temperatures. These effects bias our regional water balance over the west coast to enhanced ET/P and reduced R/P compared to observations in Fig. 2a. The inability of the CTRL simulation to capture strong horizontal gradients in R and ET also explains the underestimated spatial standard deviations over the west coast and the mountainous west. Nonetheless, the model captures the major variability of R and ET partitioning between the three regions, although there should be awareness towards the underestimated R over the west coast.The strong differences in partitioning of R and ET across the three regions imply that for a given change in the partitioning of ET and R, the relative change in R, and hence streamflows, is most sensitive over the mountainous west and the least sensitive over the west coast. Fig. 2b shows the absolute change in the partitioning of ET, R and S with respect to P with warming, highlighting the varying responses of each region. We find no noticeable changes over the west coast, in agreement with a scaling of R and streamflows to mostly follow changes in P. Over the mountainous west, we find a substantial increase in ET/P that is balanced by a similar decrease in R/P, explaining the strong relative reduction in streamflows found in Fig. 1d. Over the east, ET/P decreases while R/P increases accordingly, explaining how the diagnosed streamflow increases over the east exceed the increases in P.Attribution to upstream precipitation and snow meltWe now further characterize the physical causes for the diagnosed disconnect of precipitation and streamflow changes across CONUS, with particular focus on high-flows. In the following, we identify high-flows as days exceeding the long-term 99.7th percentile of daily streamflow (once-per-year recurrence) based on the CTRL simulation and attribute changes in high-flow frequency with warming to changes in upstream AR precipitation, non-AR precipitation, and snow melt across CONUS (see method in the section “Attribution of high-flows to upstream drivers” and Supplementary Fig. S6). We first consider maps of the long-term mean high-flow frequency in Fig. 3 and then point out the regionally varying seasonalities of high-flows and their drivers in Fig. 4.Fig. 3: Regionally varying upstream drivers of high-flows.a, c and e show fraction fdriver of high-flow events attributed to AR-precipitation events, non-AR precipitation events and snow melt events, respectively, relative to the total number of attributed high-flows based on CTRL simulation (following the attribution method described in the section “Attribution of high-flows to upstream drivers”). b, d and f, show relative changes in the respective fdriver with warming based on P2K-CTRL, normalized by the global mean warming signal. Model results are averaged over HUC4 hydrological sub-regions98, the outline of which is colored by belonging large-scale region of the west coast (green), the mountainous west (red) and the eastern US (blue).Full size imageFig. 4: Varying high-flow seasonality by region and upstream driver.a, c, e and g show regionally and monthly averaged seasonal cycles of high-flow occurrence rates in CTRL simulation, respectively for total high flow events, AR precipitation events, non-AR precipitation events and snow melt events. b, d, f and h show the respective absolute changes with warming, normalized by the global mean warming signal.Full size imageThe left column of Fig. 3 shows the fraction fdriver of high-flow events attributed to specific upstream liquid water inputs relative to the total number of attributed high-flow events. The results of the attribution are aggregated over hydrological sub-regions defined through 4-digit hydrological unit codes by the US Geological Survey (USGS)98. The separation of hydrological drivers of high streamflows follows a remarkably similar spatial pattern as streamflow changes with warming (Fig. 1d and e), indicative of the hydrological driver being a good predictor of the change in streamflow with warming. We find on the one hand that over both the West Coast and eastern US, liquid precipitation is the major driver of high-flows, with AR precipitation exceeding non-AR precipitation in most sub-regions. On the other hand, melt is found as the exclusive driver of high-flows across the mountainous west, also contributing up to 20% towards high-flow events over some sub-regions of the west coast, like the Sacramento River basin and New England in the eastern US. Note that we also assessed events of combined (AR-) precipitation and melt for cases where a single driver cannot be attributed to a high-flow. However, since we find only relatively few of such combined events (see Supplementary Fig. S7), we excluded further analysis from the scope of the paper, which focuses on the first-order regional variance within CONUS. We additionally want to point out that the results presented in this section are not sensitive to the choice of at least one different AR tracking algorithm (see the section “Atmospheric river detection and evaluation” and Supplementary Figs. S8 and S9).The right column of Fig. 3 shows the change in the number of high-flow days induced by the respective hydrological drivers with warming in our P2K experiment relative to the CTRL experiment. The pattern of these changes largely resembles the pattern of overall streamflow changes diagnosed in Fig. 1d and e. We find an increasing number of high-flow days across the west coast and the eastern US that we can now attribute to both AR precipitation and, to a slightly lower degree, non-AR precipitation. Over the mountainous west, we identify a strong reduction in melt-induced high-flows that regionally exceeds −30% K−1. This reduction in melt-driven high-flows occurs despite strong increases in precipitation over the mountainous west, particularly from ARs (Fig. 1). To start exploring why the reduction in melt-driven high-flows over the mountainous west is not buffered by increases in precipitation-driven high-flows, while precipitation-driven high-flows on the west coast and the eastern US amplify, we now take a seasonal perspective at the changes in high-flows and their drivers. This enables a more process-oriented understanding when connecting these results to changes in land-hydrological constraints on ET in the next section.Figure 4 shows the regionally and monthly integrated fraction ({f}_{{rm{driver}}}^{{rm{mon}}}) of high-flows relative to the annual sum of total attributed high-flows for the CTRL simulation (left column) and the P2K−CTRL difference normalized by the global mean warming signal (right column). The seasonality of total high-flows found in the CTRL simulation (Fig.4a) differs significantly between the three regions. Over the West Coast, high flows occur most frequently in winter months with a maximum in February. As expected for the west coast99, the attribution to upstream drivers reveals that the majority of these winter events are driven by ARs, with non-AR precipitation playing a secondary role in winter and melt becoming more significant in March. Over the mountainous West, high-flow occurrence peaks in March and is found to be exclusively driven by melt. Over the eastern US, the seasonality is less confined than in the other regions, with a broader peak around May. ARs contribute 60% of high-flows throughout spring, while non-AR events become more significant towards summer.These regional seasonalities of high-flows change under warming (right column of Fig. 4). Over the west coast, the seasonality becomes more pronounced with more frequent high-flows in winter months and less frequent high-flows in spring, driven by more frequent AR-driven high-flows and less frequent melt-driven high-flows. Over the eastern US, high-flow frequency increases throughout the year with a less pronounced maximum in spring, weakening the seasonality of high-flow occurrence with warming. Over the mountainous west, high flows diminish with warming at a rate that is proportional to the monthly climatological high-flow frequency, also implying a weakened seasonality. It is worth noting that the near-zero change in precipitation-driven high-flows at the annual mean level over the mountainous west (Fig. 3b and d) holds true for the monthly level, i.e., there are no compensating changes in precipitation-driven high-flows between different months, despite the previously diagnosed increases in mean and extreme precipitation. This motivates examination of further constraints on high-flows, such as changes in the seasonal surface water balance that we discuss in the next section.Attribution to seasonal constraints on evapotranspirationWe now explore how changes in energy and water constraints on ET with warming help explain changes in the regional variability in changes of the partitioning between ET and R, as identified in Fig. 2b, and hence the warming response of high-flows. We start by considering annual mean maps of the annual mean dryness index ({mathcal{D}}) and evaporative index ({mathcal{E}}) (as commonly used in Budyko analysis100,101):$${mathcal{D}}={{PET}}/P$$
    (2)
    $${mathcal{E}}={{ET}}/P$$
    (3)
    where PET refers to potential ET, i.e., the energetic upper bound on ET (see the “Methods” subsection “Calculation of potential evapotranspiration”). An annual mean ({mathcal{D}} < 1) indicates that ET is energy limited, while ({mathcal{D}} > 1) indicates that ET is water limited. Figure 5a shows that annual mean ({{mathcal{D}}}_{{rm {CTRL}}}) follows a somewhat similar spatial pattern as the streamflow response to warming (Fig. 1d and e), with a strong zonal gradient between energy-limited ET in the eastern US and some parts along the west coast, and a water-limited mountainous west with pronounced aridity in the south-west. This general pattern is in agreement with downscaled reference datasets102,103,104.Fig. 5: Energy and water constraints on changes in seasonality of surface water balance.a Shows annual mean map of ({mathcal{D}}) in CTRL simulation, where PET is derived from Penman–Monteith equation over an open water surface (see the “Methods” subsection “Calculation of potential evapotranspiration”). ({mathcal{D}} <)1 indicates energy limitation of ET, and ({mathcal{D}} >)1 indicates water limitation of ET. b Shows annual mean change in ({mathcal{E}}) with warming, indicating whether ET or P is more enhanced with warming. c–h Show relative changes of mean seasonal cycle c in P, d in ET, e in ({{mathcal{E}}}_{{rm {m}}}) (Eq. (5)), f in ({{mathcal{D}}}_{{rm {m}}}) (Eq. (4)), g in Fnet, normalized by specific heat of vaporization ({{mathcal{H}}}_{{rm {vap}}}) and ETCTRL, h in R. Relative changes are calculated after spatial averaging.Full size imageWith an energetic surplus in evaporative potential across the mountainous west, ET is found to increase at a higher rate than over the west coast or eastern US, as implied by a positive change in ({mathcal{E}}) shown in Fig 5b. As previously discussed in the section “Precipitation and streamflow changes with warming”, the increase in ({mathcal{E}}) across the mountainous west implies a shift of the surface water balance towards overall drier conditions with reduced R. Over the eastern US a slight opposite shift towards enhanced R is found and over the west coast ({mathcal{E}}) changes are only slightly positive. These energetic constraints on annual mean ET align well with the diagnosed annual mean changes in streamflows in Fig. 1.To more closely examine how the discussed annual mean changes in ({mathcal{D}}) and ({mathcal{E}}) occur in different regions, we now turn to the seasonal perspective and first consider changes in ET and P (Fig. 5c and d). All regions exhibit a pronounced seasonality in P changes with warming, where winter P increases at a higher rate than summer P. This seasonality is particularly pronounced over the mountainous west, where mean P increases at around 6% K−1 in winter and reduces in July and August at around −1% K−1. The seasonality in changes of ET (Fig. 5d) is similarly phased as P; however, the relative increase in ET with warming is about double the relative increase in P over the mountainous west and also around one third higher over the west coast in winter, while it is slightly weaker over the east.To consider monthly changes in ({mathcal{D}}) and ({mathcal{E}}), we define them to more appropriately represent the water available for ET on the monthly timescale, which is rather the sum of liquid precipitation Pl and melt M than total precipitation P:$${{mathcal{D}}}_{{rm {m}}}={{PET}}/({P}_{{rm {l}}}+M)$$
    (4)
    $${{mathcal{E}}}_{{rm {m}}}={{ET}}/({P}_{{rm {l}}}+M)$$
    (5)
    This definition excludes frozen precipitation that does not melt throughout the month and includes melt that stems from frozen water storage that may have formed through snowfall in months prior.We find that despite the discussed strong increases of ET in winter, ({{mathcal{E}}}_{{rm {m}}}) and ({{mathcal{D}}}_{{rm {m}}}) (Fig. 5e and f) reduce in winter months in the eastern US and the mountainous west at up to −20% K−1 while increasing in spring. This counter-intuitive result is explained by seasonal changes to the liquid water supply Pl + M, which outweigh the strong changes in ET (see Supplementary Fig. S10 and decomposition in Supplementary Fig. S11). During winter months, liquid water supply for ET is enhanced with warming as precipitation shifts from frozen to liquid and snow melts throughout the season, shifting the winter-time mountainous west from a water-limited ET regime to an energy-limited one (triangles in Fig. 5f). During spring, reduced snow-cover reduces melt and hence liquid water supply, increasing ({{mathcal{E}}}_{{rm {m}}}) and ({{mathcal{D}}}_{{rm {m}}}).The hydrological regime shift of the winter-time mountainous west towards an energy-limitation of ET occurs despite an enhanced energy supply in winter, as shown by the relative change in the net surface energy imbalance Fnet (Fig. 5g), which we normalize by the latent heat of vaporization ({{mathcal{H}}}_{{rm {vap}}}) and ETCTRL to obtain comparability to ET changes. The increased Fnet in winter can largely be attributed to the shortwave component, which is strongly controlled by changes in surface albedo through the loss of reflective snow (Supplementary Figs. S10 and S12). This finding is in agreement with previous studies over the western US that argue towards loss of reflective snow energetically fueling ET under warming91,105. Our results highlight how this increased energy-supply for ET is present but dominated by an enhanced water-supply that alters the winter-time mountainous west to an energy-limited ET regime.Relative changes in winter-time Fnet in the other regions are weaker than over the mountainous west (Fig. 5g) as the energy balance of the surface is less dominated by snow. In fact, over the eastern US most of the increase in Fnet is explained through the longwave component, which is mostly driven by the water vapor feedback as cloud changes are small and towards reduced cloud cover in all regions (Supplementary Figs. S10 and S12). Overall, these less pronounced changes to Fnet under an already existing energy limitation on ET in the CTRL simulation yield overall weaker responses of ET to warming compared to the mountainous west.Finally, having discussed the seasonal changes in P and ET with warming across CONUS, we turn to runoff R (Fig. 5h), the direct precursor of streamflows. R changes exhibit a strong seasonality over the mountainous west with increases of around 10% K−1 in winter and summer, and a strong reduction near 20% K−1 in spring. Considering that in absolute terms most R is produced in spring, we find a significant reduction of the annual mean R of −4.1% K−1 (see abs. changes in Supplementary Fig. S10). The increase in winter- and summer-time R is nonetheless worth noting, given that in both seasons ET is found to increase at a higher rate than P (Fig. 5c and d). In winter months, the increased supply of liquid P and melt explains increases in both R and ET (Supplementary Fig. S10). The increase of R in summer despite a reduction in mean P (Fig. 5c) is most likely explained by a shift towards less frequent but more intense P events. Over the west coast, R changes also have a clear seasonal component, albeit at a lower magnitude than over the mountainous west, with increases at up to 5% K−1 in winter. Over the Eastern US, R increases between 5% and 10% K−1 throughout the year with a slight variation between a weaker increase in March due to loss of melt and a stronger increase in May due to increased liquid precipitation, particularly from ARs (Fig. 4). The stronger increase in R over the east compared to the west coast can be attributed to the overall stronger increase in P over the east under similar changes in ET, yielding a change in partitioning of P towards increased R/P and reduced ET/P as shown in Fig. 2b.Relative importance of changes in antecedent upstream soil moisture and water inputsWith insights gained into the seasonal character of the processes driving changes in P, ET and R across CONUS, we now turn back to changes in high-flows and discuss how these characteristics tie in to changes in antecedent soil moisture, which is often regarded as a key predictor for R production31,32,33,34. Specifically, we examine shifts in antecedent soil saturation ({{mathcal{S}}}_{{rm {a}}}) of the top 1 cm of the soil, which describes the degree to which the pores of the soil’s top layer are filled with water prior to a precipitation or melt event. The upper-most layer is typically considered for this purpose because it controls production of R at the surface, which during an extreme precipitation or melt event is key for total R production31. Our approach is to take advantage of the day-by-day comparability between our CTRL and P2K simulations that is achieved through tightly nudged atmospheric winds, and allows us to separate identified high-flow days into those that occur only in the CTRL simulation, only in the P2K simulation or in both. This way, we can ask across these three categories of events how a change in upstream ({{mathcal{S}}}_{rm {{a}}}) conditions may have contributed to a high-flow being only present in either CTRL or P2K, and put this change in relation to changes in the antecedent upstream soil-top water inputs ({{mathcal{P}}}_{{rm {a}}}), i.e. the sum of antecedent upstream precipitation and melt (detailed the “Methods” subsection “Quantifying antecedent soil conditions and upstream drivers ”).The bottom panels of Fig. 6 depict how many high-flow events, identified on a specific day and location, fall into the different categories of whether they occurred in either only the CTRL or the P2K simulation or in both, showing the total number and its breakdown into the various upstream drivers for each of the CONUS sub-regions. For each of the three categories, the upper panels of Fig. 6 display box-plots for distributions of antecedent upstream conditions (({{mathcal{S}}}_{{rm {a}}}) and ({{mathcal{P}}}_{{rm {a}}})) for all high-flow events in the CTRL and P2K simulations. It is evident that the dominant driver of high-flow events and its dependence on the climate state vary strongly across the sub-regions, as would be expected from our previous analysis of Fig. 3. Over the west coast, there are few high-flows that only occurred in the CTRL simulation, while most high-flows occur in both CTRL and P2K and are mostly driven by AR precipitation. Those high-flows that only occur in the P2K simulation show slight increases in ({{mathcal{P}}}_{{rm {a}}}) and ({{mathcal{S}}}_{{rm {a}}}) with warming, both of which act towards enhancing streamflows. As diagnosed in Fig. 4b, the additional high-flows over the west coast in the warmer climate occur mostly in winter during peak AR season, in which enhanced liquid precipitation and earlier snow-melt moisten soils, creating favorable conditions for enhanced runoff.Fig. 6: Comparing magnitudes of antecedent soil moisture and upstream driver changes of high-flows.Comparison of changes in antecedent soil moisture and upstream drivers for a) west coast, b) mountainous west and c) eastern US. For each sub-plot, the bottom panels show the regionally integrated distributions of high-flows identified in either CTRL or P2K simulation in categories of whether each high-flow, identified on a specific day and location, only occurred in the CTRL simulation, the P2K simulation or in both. The total number for each category is broken down into the attributed upstream drivers (AR precipitation, non-AR precipitation, melt). Upper panels show box-plots (black line at median, box-bounds at 25th and 75th percentiles, whiskers at 10th and 90th percentiles) of the total distribution (i.e. all upstream drivers) of upstream antecedent soil saturation ({{mathcal{S}}}_{{rm {a}}}) (left y-axis) for high-flow events in the CTRL (orange) and the P2K (red) simulations, as well as sum of all upstream drivers ({{mathcal{P}}}_{{rm {a}}}) (right y-axis) for high-flow events in the CTRL (light blue) and the P2K (dark blue) simulations in each category.Full size imageThe strong reduction of high-flows over the mountainous west with warming (Fig. 3) is reflected in the bottom panel of Fig. 6b showing that most high-flows occur only in the CTRL simulation and are induced by melt. Looking at the difference in ({{mathcal{S}}}_{{rm {a}}}) and ({{mathcal{P}}}_{{rm {a}}}) for these cases between CTRL and P2K, we find that ({{mathcal{P}}}_{{rm {a}}}) reduces significantly in the warmer climate while ({{mathcal{S}}}_{{rm {a}}}) increases. Hence, the reason many high-flows over the mountainous west in the CTRL simulation do not exist in the P2K simulation is because of the loss of upstream water inputs ({{mathcal{P}}}_{{rm {a}}}), in particular snow-melt, which outweighs the increase in increased ({{mathcal{S}}}_{{rm {a}}}) that by itself favors high-flows in the P2K simulation. Those high-flows over the mountainous west that are only present in the P2K simulation are samples from a much more even distribution of precipitation and melt-induced cases (Fig. 6b, bottom panel), indicating a shift towards a more mixed and less snow-dominated hydrological regime. To a large extent, the newly arising high-flows in the P2K simulation are enabled by significantly enhanced ({{mathcal{S}}}_{{rm {a}}}), which on the median increases from 0.2 in CTRL to 0.45 in P2K while the upstream drivers ({{mathcal{P}}}_{{rm {a}}}) show a less significant increase. This strong increase in ({{mathcal{S}}}_{{rm {a}}}) can be understood as a direct consequence of the discussed reduction in dryness index ({mathcal{D}}) (Fig. 5f) in winter months over the mountainous west with warming.Over the eastern US, the largest number of high-flows occurs only in the P2K simulation, highlighting the strong sensitivity to warming in this region, which is driven to a similar degree by AR and non-AR precipitation. The high-flows that arise only in the P2K simulation are driven by a shift and a narrowing of the ({{mathcal{S}}}_{{rm {a}}}) distribution towards higher ({{mathcal{S}}}_{{rm {a}}}) values, as well as more extreme ({{mathcal{P}}}_{{rm {a}}}), similar to the West Coast, but to a more pronounced degree. The enhanced sensitivity of ({{mathcal{S}}}_{{rm {a}}}) over the eastern US compared to the west coast is likely caused by the fact that over the west coast, ({{mathcal{S}}}_{{rm {a}}}) is already close to saturation in the CTRL simulation for most high-flows, while over the eastern US ({{mathcal{S}}}_{{rm {a}}}) increases in winter months as the dryness index ({{mathcal{D}}}_{{rm {m}}}) reduces significantly (Fig. 5f).DiscussionWe conducted a counter-factual warming experiment with the GFDL moderately high-resolution coupled atmosphere-land model AM4.0/LM4.0 to derive a physical understanding of the relation of the precipitation and streamflow responses to warming across CONUS, with particular focus on high-flows that are indicative of flooding. By nudging atmospheric winds in both our control and warming simulations, our approach isolates the thermodynamic component of the total response, focusing attention of this study on the implications of the first-order and most well-understood part of the total precipitation response on streamflows. We first quantified the precipitation and streamflow responses on the mean and the extremes across CONUS, elucidating their regionally varying relationship through the varying partitioning of ET and R. We then attributed the relation of precipitation and streamflow responses to changes in upstream water inputs, distinguishing AR and non-AR precipitation as well as melt, seasonal energy and water constraints on ET and associated implications for antecedent soil moisture conditions prior to high-flows.We identify three sub-regions of CONUS that exhibit substantially varying streamflow responses to a comparatively homogeneous precipitation response. Over the West Coast, the streamflow response is positive but falls below precipitation scaling, approaching only up to 4% K−1 in the extremes. Over the Eastern US, streamflow scales beyond precipitation, approaching 10% K−1 and over the mountainous western US, streamflows dwindle, approaching −10% K−1 despite significant increases in annual mean precipitation, particularly from ARs. We associate these regionally varying relationships of the thermodynamic precipitation and streamflow responses to varying hydrological regimes in terms of dominance of snow compared to liquid precipitation, as well as energy and water limitations of ET, which are embedded in regionally distinct seasonalities.The most striking dichotomy between dwindling streamflows across the mountainous west and increasing streamflows elsewhere can be attributed to the dominance of springtime snow melt in driving the regional hydrological dynamics over the mountainous west. Reduced accumulation of snow throughout the winter months diminishes melt in spring, the season of most significant runoff production. As a consequence of reduced snow-cover in winter, the surface energy balance shifts towards higher PET through enhanced short-wave absorption. At the same time, liquid water supply increases with enhanced winter-time liquid precipitation and melt, yielding an overall moistening of the mountainous west in winter, fueling ET increases at about double the rate of the mean precipitation response and also in winter-time runoff. Through these changes, the winter-time mountainous west shifts from a water-limited ET regime to an energy-limited one with warming. We find that the moister soils not only enable enhanced ET, but also create favorable conditions for high-flows during winter at a more even distribution among driving AR precipitation, non-AR precipitation and melt. However, in the annual mean, these increases do not balance the significant reduction of melt-driven runoff in spring, resulting in a net annual runoff loss of around −4.1% K−1.While both the eastern US and the west coast exhibit increasing streamflows with warming, the increase is significantly stronger across the East. Over both regions, liquid precipitation is identified as the main driver for high-flows, with ARs contributing about two-thirds over the west coast and about half over the east, while melt plays a secondary role. These proportions are insensitive to the use of a different AR tracking algorithm, enhancing confidence in the significant role ARs play in thermodynamically scaling as a strong contributor to high-flows with warming. However, we need to recognize that other types of precipitation systems, such as meso-scale convective systems or hurricanes, have not been assessed in this study, and possible misrepresentations of those could bias the partitioning towards the relatively well-captured ARs. Expanding the attribution framework of this study to more types of precipitation systems denotes an interesting future perspective.To explain the diagnosed difference in the magnitude of the streamflow increase between the West Coast and the eastern US, our results suggest two major reasons. Firstly, the partitioning of water losses between ET and R vary significantly between the two regions, with R/P at 51 % over the west coast and about 28 % over the eastern US. This difference implies a stronger sensitivity of relative changes in R to changes in the partitioning of ET and P to warming over the east. Secondly, the precipitation increase over the east is simply at a higher magnitude than over the west (Fig. 5c), which in conjunction with almost identical increases in ET must yield a stronger runoff response over the east. This finding is in agreement with previous studies that identify the north-eastern US as the region with the highest precipitation increase over the last decades across the US67,106,107.This work highlights the complex system of regionally confined hydrological processes, which result in strongly varying streamflow sensitivities to warming. Changes in upstream water inputs shift the prediction problem of high-flows towards one of precipitation and away from melt, particularly over mountainous regions. ARs are identified as the strongest driver of high-flows across CONUS, not just the West Coast, and react sensitively to warming, implying a need for improving the predictability of ARs in particular. The presented work isolates the thermodynamic component of the total warming response, which, for precipitation extremes, is thought to dominate the more uncertain dynamic response, but may not capture significant contributions towards the regional streamflow response. Future research should focus on expanding the framework presented here towards incrementally including the dynamical response, the direct effect of CO2 and its feedback with vegetation, aerosol forcing, as well as changes in SST patterns.Land surface modelling advances to further reduce approximations of sub-grid heterogeneity and more sophisticated representations of river streamflow dynamics will help reduce uncertainties and more explicitly represent flooding. Another avenue of ongoing development is global storm resolving models (GSRMs) that run at grid-spacings of 4 km and less108,109. While studies of land–atmosphere interactions based on such models are just starting, first results point to significant discrepancies with moderately high-resolution models with regard to the sensitivity of snow-cover to warming in mountainous regions110. Further exploring these differences, also in light of the currently simplified assumptions made in the land schemes of GSRMs, denotes an exciting future perspective.MethodsModel setup and simulationsWe use the coupled land–atmosphere model AM4.0/LM4.0 developed at NOAA’s Geophysical Fluid Dynamics Laboratory (GFDL), the main characteristics of which are described in the reference papers46,58,59. The component models AM4.0 and LM4.0 are extensively used and validated in studying various aspects of the earth system, as they denote integral components of GFDL’s modeling suite, namely SPEAR (Seamless System for Prediction and EArth System Research)111 and CM4 (Climate Model version 4)112, as well as the physical baseline for more comprehensive representations of dust and chemical interactions within ESM4 (Earth System Model version 4)113,114. A major simplification to those comprehensive modeling systems is that our AM4.0/LM4.0 simulations are forced by observed SSTs and sea ice concentrations, often referred to as AMIP (Atmospheric Model Intercomparison Project) mode115, which omits uncertainties revolving around ocean-atmosphere interactions that can result in biases of SST patterns.We run AM4.0/LM4.0 in a moderately high resolution mode at around 50 km grid spacing, which on the model’s cubic-sphere grid is represented by 192 × 192 grid-boxes per cube face. This version of the model contributed to the CMIP6 HighResMip experiments37 and our simulations follow the HighResMip protocol in being forced by daily varying observed SSTs, sea-ice concentrations and radiative gases. One major feature of our simulations is that we nudge atmospheric winds against the 6-hourly NCEP reanalysis66 to achieve day-by-day comparability to observed weather and between our control and warming simulations, isolating the thermodynamic warming response. We conduct the nudging with a tight nudging timescale of 30 minutes, which assures that there is no significant change in circulation patterns between our control and warming simulations (Supplementary Fig. S4). The nudging is applied at a 30 min time scale by linearly interpolating the 6-hourly reanalysis data in time to the model timesteps and applying a tendency term to the wind-fields that is proportional to the interpolated NCEP winds and an exponential with an e-folding time of 30 min.Our control (CTRL) and warming (P2K) simulations only differ in forcing by a homogeneous increase of daily SSTs by 2 K. A realistic future warming pattern can be decomposed into global mean and SST warming patterns. The goal of this study is to focus on the global mean SST warming component, which is robust in both models and observations, although the magnitude of global mean warming in response to a fixed radiative forcing remains uncertain116. Changes in SST warming patterns are challenging to capture in coupled climate models117,118, implying uncertainties for precipitation changes with warming patterns119,120 that we exclude from the scope of this study.The land surface model LM4.0 discharges surface and sub-surface runoff across a grid-scale river network as fluxes of liquid water, ice and heat to connected grid-cells following the hydraulic geometry approach60,61. This representation of rivers is idealized by assuming 1-way fluxes with the surrounding land and neglecting the effects of inundation. However, for the scope of this study to obtain a regional understanding of mean and high-flow responses to warming, this representation is adequate, given previous evaluation of the model for annual mean runoff46,121, runoff ratios61, and monthly mean peak discharge in the Mississippi122.Atmospheric river detection and evaluationFor AR detection, we follow the method of Guan and Waliser123, which has previously been applied to AM4.0/LM4.0 simulations, with good agreement to observations45,46. The algorithm is based on 6-hourly mean outputs of zonal and meridional vertically integrated vapor transport (IVT) to compute IVT magnitude at each grid-cell. The climatological seasonal 85th percentile of IVT (considering all 6-hourly timesteps within 5-month intervals centered around respective months) is calculated at each grid-cell and used as a threshold to identify possible AR objects, together with a lower absolute limit for IVT of 100 kg m s−1. Following previous studies46,124, we calculate separate IVT thresholds (i.e., 85th percentile of IVT) for the CTRL and the P2K simulation to eliminate changes in AR frequency purely due to Clausius-Clapeyron scaling of water vapor content in the atmosphere, which in itself does not imply enhanced AR activity or impacts. Changes in AR frequency with warming are shown in Supplementary Fig. S15, showing small changes of less than 1% K−1 over the west coast and eastern US, but significant increases over the mountainous west. This regional dependence highlights that our tracking method is not subject to a strong sensitivity simply due to increases in water vapor following Clausius–Clapeyron.After applying the IVT threshold, geometric conditions are applied to the AR candidate objects (i.e., spatial envelopes of regions of sufficient IVT). The geometric conditions include a minimum AR length > 2000 km, a length-to-width ratio > 2, as well as directional requirements with a minimum poleward component of IVT > 50 kg m s−1 and a coherence of IVT direction with more than 50% of AR-candidate pixels required to have IVT directions within 45° of the mean AR-candidate object direction. The identified AR-objects are stored on a 6-hourly timestep. Since for the rest of our analysis we use daily data, we map the 6-hourly AR-masks to daily masks by considering a given day an AR-day if an AR is identified at least once on any of the 6-hourly timesteps. More detailed descriptions of how geometric requirements are implemented (i.e., through definition of an axis) can be found in the paper introducing the original algorithm123. An additional criterion we apply for the AR condition is a minimum precipitation of 1 mm day−1 (following ref. 46).We evaluated the resulting spatial distribution of AR frequency (Supplementary Fig. S14) and the AR contribution to mean and extreme precipitation across CONUS against observations (Supplementary Fig. S13). Both, the distributions of AR frequency and the climatological contribution of AR precipitation to annual mean and extreme precipitation is found to be very similar between our CTRL simulation and observations across CONUS.Given uncertainties among AR tracking methods125, we validated our results in the section “Attribution to upstream precipitation and snow melt” on ARs as upstream drivers of high-flows by reproducing them based on another AR tracking algorithm126. A key difference of this algorithm is that it first calculates IVT anomalies by subtracting the climatological day-of-year mean IVT at each grid-cell and then uses an IVT-anomaly threshold rather than a threshold on absolute IVT (as opposed to ref. 123). We calculate the IVT-anomaly threshold as the 94th percentile of 6-hourly anomalies from the day-of-year mean following previous studies126,127,128,129. In addition, we require a minimum precipitation of 1 mm day−1, as we do with the first AR-tracking algorithm. Further details about geometric requirements can be deduced from ref. 126.We find varying differences in AR frequency between the two algorithms depending on the region (Supplementary Fig. S14), finding a lower AR frequency over the mountainous west (as percentage of time: 5.8% based on123 and 2.7% based on ref. 126) and the west coast (11.3% based on ref. 123 and 10.1% based on ref. 126) and a slightly higher AR frequency over the eastern US (12.5% based on ref. 123 and 12.6% based on ref. 126). Despite these slight regional differences, the reproduced Fig. 3 and Fig. 4 based on the second AR tracking algorithm (Supplementary Figs. S8 and S9) look remarkably similar to the results based on the first algorithm, enhancing confidence in our conclusions.Evaluation of the surface water balanceTo evaluate the regional surface water balance of our AM4.0/LM4.0 CTRL simulation, we make use of an observationally deduced 1 km resolution water balance dataset available across CONUS from 1980 to present95,96. The dataset is based on a Thornthwaite-style water balance model94, which takes observed daily precipitation and temperature data from Daymet Daily data version 397 and computes daily runoff R, storage changes (soil water (Delta {{mathcal{S}}}_{{rm {tot}}}) and snow water equivalent ΔSWE) and ET. The water balance model is detailed by Terceck et al. 95 and the data and code are freely available at NASA-Earthdata (https://cmr.earthdata.nasa.gov/search/concepts/C2674694066-LPCLOUD.html, last access: 07/27/2025).To make the reference data comparable to our model output, we bi-linearly interpolate the dataset to the model grid and select the period 1980–2020. We first compute monthly averages of the fluxes the dataset provides, namely liquid precipitation Pl, ET, and R. Since frozen precipitation is not provided explicitly, we compute it as the residual of the monthly mean surface water balance:$${P}_{{{f}}}={{ET}}+R+Delta {{mathcal{S}}}_{{rm {tot}}}+Delta SWE-{P}_{{rm {l}}}$$
    (6)
    We then obtain total precipitation as P = Pl + Pf and deduce the regional, annual averages and spatial standard deviations across the west coast, mountainous west and eastern US in the same way as for the CTRL simulation and the results are presented in Fig. 2a. In addition, climatological maps comparing the ET/P and R/P across CONUS between the CTRL simulation and the reference dataset are provided in Supplementary Fig. S5.Attribution of high-flows to upstream driversWe identify daily high-flows from the output of the coupled grid-scale river network that is part of LM4.0 (Supplementary Fig. S1). We do so by defining a streamflow threshold based on the long-term 99.7th percentile of daily streamflows at each grid-point in the control simulation, which represents a recurrence interval of once per year. By applying this threshold to both CTRL and P2K simulations, we obtain a database of high-flow days at each location in both simulations. For each of these high-flow days, we conduct an attribution analysis to identify the major upstream source of water that led to the high-flow, distinguishing AR precipitation, non-AR precipitation and snow melt. For a given event, we first identify the upstream area, which for each grid-point is defined by the grid-scale river network. We then calculate the upstream area integral of the potential high-flow drivers, so we obtain their upstream time-series. We then conduct a lagged temporal correlation analysis between downstream streamflow at the grid-point where the high-flow occurred and the area-integrated upstream drivers.To conduct the lagged temporal correlation analysis, we first calculate an event-specific drainage timescale τdrain, which estimates the time it takes for water entering the upstream catchment to drain into the downstream grid-point:$${tau }_{{rm {drain}}}=frac{{v}_{{rm {ds}}}}{sqrt{A/pi }}$$
    (7)
    where vds is the streamflow velocity at the downstream grid-point on the day of the high-flow and A is the upstream area of the catchment. Hence, τdrain represents the time it would take for water in the center of a round catchment with area A to escape the catchment on a straight line at a fixed streamflow velocity vds. Although this assumption is very crude for estimating the actual drainage time of precipitation entering the catchment, we use it as a starting point for identifying a sufficient correlation of upstream drivers and downstream streamflows.Knowing τdrain for a given high-flow event, we now iteratively determine the best-correlated upstream driver. We do so by shifting the timeseries of upstream-integrated drivers by τshift ∈ (0, 1, …, 2 × τdrain) days. For each iteration, we select a temporal window of downstream streamflow and upstream drivers of ±τdrain around the day of high-flow. As a first check of whether a potential upstream driver could have caused the high-flow, we check whether the temporal integral over the selected temporal window of an upstream driver is larger than the temporal integral of streamflow. For example, in the selected window there must have been more precipitation upstream than discharge downstream, if precipitation is to be attributed the main driver of the discharge. We conduct this test for each τshift and each of the upstream drivers AR-precipitation, non-AR precipitation and melt as well as the sums of AR-precipitation and melt and non-AR precipitation and melt. If for any of these cases, the test is passed, we calculate the temporal correlation coefficient ρD,S of upstream driver and downstream streamflow in the selected temporal window. If ρD,S > 0.8, we attribute the high-flow event to the particular driver. Note the cases of combined AR- or non-AR precipitation with melt occurred so infrequently in our analysis that we omitted them in our discussions.A schematic depiction of the attribution method’s workflow and some evaluation metrics are provided in Supplementary Fig. S6. The method is able to attribute about 35% of high-flows in both the CTRL and the P2K simulation, and works best for grid-points with small upstream basins, which is expected given the idealized assumptions about shape and drainage dynamics of the upstream basin as outlined above through the attribution method.Quantifying antecedent soil conditions and upstream driversThe characterization of antecedent soil saturation is embedded in our attribution method of upstream high-flow drivers (see Supplementary Fig. S16). For each high-flow event, we store the timeseries of the area sum of soil moisture at 1 cm depth over the upstream basin, which we call qs, within the temporal window of ±τdrain around the respective high-flow days. Using Eq. (8), we transform qs into volumetric units, referred to as qv, which is divided by soil porosity φ (Eq. (9)) yields soil saturation in 1 cm depth ({mathcal{S}}). Following Eq. (10), we calculate the antecedent upstream soil saturation as the minimum ({mathcal{S}}) within the period leading up to the high-flow event constrained by τdrain.$${q}_{{rm {v}}}=frac{{q}_{{rm {s}}}}{1000,{rm{kg}},{{rm{m}}}^{-2}}$$
    (8)
    $${mathcal{S}}=frac{{q}_{{rm {v}}}}{varphi }$$
    (9)
    $${{mathcal{S}}}_{{rm {a}}}=min ({mathcal{S}}(t))quad {rm{for}},{t}_{{rm {HF}}}-{tau }_{{rm {drain}}}le t < {t}_{{rm {HF}}}$$
    (10)
    To compare ({{mathcal{S}}}_{{rm {a}}}) between CTRL and P2K for events that only occur in CTRL and not in P2K or the reverse, we apply our high-flow attribution method based on the CTRL high-flow mask (i.e., identified days and locations of high-flows in CTRL) to the P2K simulation, as well as the P2K high-flow mask to the CTRL simulation. This way, we can obtain ({{mathcal{S}}}_{{rm {a}}}) for the CTRL and P2K scenario for all events that denote a high-flow only in CTRL, only in P2K or in both.The sum of upstream drivers ({{mathcal{P}}}_{{rm {a}}}), made up of AR precipitation PAR, non-AR precipitation PnAR and melt M, depicted in Fig. 6, is derived analogously to Sa in that we consider the timeseries of the area-sum of upstream drivers leading up to the high-flow event. However, in this case, we consider the temporal integral of all drivers following Eq. (11).$${{mathcal{P}}}_{{rm {a}}}=mathop{int}nolimits_{t-{tau }_{{rm {HF}}}}^{{t}_{{rm {HF}}}}({P}_{{rm {AR}}}+{P}_{{rm {nAR}}}+M),{rm {d}}t$$
    (11)
    Calculation of potential evapotranspirationPotential evapotranspiration (PET) can be considered an energetic upper bound of how much a surface can evaporate given its energy imbalance and in some definitions given specific atmospheric conditions. A wide variety of approaches to calculate PET exist, which are designed for specific vs. more broad applications and that have varying requirements towards the data130,131. Here, we use the widely used Penman–Monteith Equation (Eq. (12)) over an open water surface131 to conduct our analysis of ET changes with warming in the section “Attribution to seasonal constraints on evapotranspiration”.$${{PET}}=frac{Delta }{Delta +gamma }frac{({F}_{{rm {net}}}-G)}{{L}_{{rm {v}}}}+frac{gamma }{Delta +gamma }frac{6.43(1+0.536u)({e}_{{rm {s}}}-{e}_{{rm {a}}})}{{L}_{{rm {v}}}}$$
    (12)
    where Fnet is the radiative surface net energy imbalance, G is the sensible heatflux into the subsurface, both (in units of MW m−2); ea and es are the 2 m actual vapor pressure and saturation vapor pressure (kPa); Δ is the derivative of es with respect to temperature T (kPa K−1); u is the windspeed in 2 m height (m/s); Lv is the latent heat of vaporization (MJ kg−1); γ is the psychometric constant (kPa K−1). These metrics are calculated through the following equations:$${L}_{{rm {v}}}=2.501-0.002361T$$
    (13)
    $${e}_{{rm {s}}}=0.6108,exp (17.27T/(T+237.3))$$
    (14)
    $$Delta =4098{e}_{{rm {s}}}/{(T+237.3)}^{2}$$
    (15)
    $$gamma =0.0016286P/{L}_{{rm {v}}}$$
    (16)
    $$u=0.75 {u}_{10{rm{m}}}$$
    (17)
    where T is 2 m temperature (°C) and P is atmospheric surface pressure (kPa) and u10m is the windspeed at 10 m height, which is the reference height we obtain as output from the model.

    Data availability

    The GFDL AM4 model source code can be obtained from https://data1.gfdl.noaa.gov/nomads/forms/am4.0/ NCEP reanalysis data used for nudging atmospheric winds is available at https://psl.noaa.gov/data/gridded/data.ncep.reanalysis.html Observationally derived CONUS water balance model output is available at https://cmr.earthdata.nasa.gov/search/concepts/C2674694066-LPCLOUD.html ERA5 and IMERG precipitation data used for evaluation of atmospheric rivers and precipitation is available at the Copernicus Climate Change Service ClimateData Store (CDS; https://cds.climate.copernicus.eu/) and at https://gpm.nasa.gov/data/directory, respectively. Observationally derived reference data for the surface water balance over CONUS is available at https://cmr.earthdata.nasa.gov/search/concepts/C2674694066-LPCLOUD.html .
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    Download referencesAcknowledgementsThis report was prepared by Marc Prange under awards NA22OAR4050663D and NA23OAR4050431I from the National Oceanic and Atmospheric Administration, U.S. Department of Commerce. The statements, findings, conclusions, and recommendations are those of the author(s) and do not necessarily reflect the views of the National Oceanic and Atmospheric Administration or the U.S. Department of Commerce. The authors want to thank Minki Hong for useful discussions about the surface hydrologicalaspects of this study. We also thank Zhihong Tan and Tom Delworth for reading and providing feedback on this manuscript, as well as Joseph Clark for providing code and assistance in setting up the Atmospheric River tracking algorithm.Author informationAuthors and AffiliationsAtmospheric and Oceanic Sciences, Princeton University, Princeton, NJ, USAMarc PrangeGeophysical Fluid Dynamics Laboratory, NOAA, Princeton, NJ, USAMing Zhao, Elena Shevliakova & Sergey MalyshevAuthorsMarc PrangeView author publicationsSearch author on:PubMed Google ScholarMing ZhaoView author publicationsSearch author on:PubMed Google ScholarElena ShevliakovaView author publicationsSearch author on:PubMed Google ScholarSergey MalyshevView author publicationsSearch author on:PubMed Google ScholarContributionsM.P., M.Z., and E.S. conceptualized the main outline of the scientific work. S.M. helped with the technical analysis of the model output. M.P. wrote the manuscript and conducted the data analysis. All authors reviewed the manuscript.Corresponding authorCorrespondence to
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    Sea level rise and flooding of hazardous sites in marginalized communities across the United States

    AbstractSea level rise (SLR) increases the risk of flooding at coastal sites that use and produce hazardous substances. We assess whether socially marginalized populations in the United States are more likely to be impacted by projected SLR-related flooding of hazardous sites that could result in contaminant releases. We identify 5500 facilities at risk of a 1-in-100-year flood event by 2100 under a scenario of continued high greenhouse gas emissions, including coastal power plants, sewage treatment facilities, fossil fuel infrastructure, industrial facilities, and formerly used defense sites. Seven states (Louisiana, Florida, New Jersey, Texas, California, New York, and Massachusetts) account for nearly 80% of projected at-risk facilities. Controlling for population density and county, a one standard deviation increase in the proportion of linguistically isolated households, neighborhood residents identifying as Hispanic, households with incomes below twice the federal poverty line, households without a vehicle, non-voters, and renters is associated with 19-41% higher likelihood of having a site at risk of SLR-related flooding within 1 kilometer (odds ratios [95% confidence intervals]: 1.19 [1.09, 1.31], 1.22 [1.08, 1.37], 1.27 [1.16, 1.39], 1.35 [1.21-1.51], 1.36 [1.21, 1.53], and 1.41 [1.32, 1.52], respectively). Results elucidate the need for disaster planning, land-use decision-making, as well as mitigation strategies that address the inequitable hazards and potential health threats posed by SLR.

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    IntroductionGlobal sea level has risen more than 11 cm over the last three decades and that rate is accelerating1, leading to an increase in coastal flooding due to high tides, waves, storm surge, El Niño events and other factors. Extreme coastal flooding is projected to more than double by 2050 across much of the world2. By 2100, nearly all of the coastal United States (U.S.) is expected to experience elevated water levels on a daily basis that today occur only twice per century3, with a rapid increase in the frequency of high tide flooding projected to begin in multiple cities during the next decade4.Extreme flood events result in the release of toxic substances into the environment. For example, over 200 contaminant releases were reported in the Texas Gulf Coast after flooding resulting from Hurricane Harvey in 2017. Over 10 million pounds of regulated air pollutants were released from refineries, petrochemical, and other industrial facilities5, and the catastrophic explosion of a chemical plant due to the loss of power for refrigeration necessitated the evacuation of 40,000 people6. Around the world, industrial facilities are disproportionately located along coastlines due to the historical importance of maritime trade to the establishment of industrial port cities, strategic access to global trade routes for raw materials and finished products via ports, and need for sea water for cooling and wastewater disposal. Marginalized racial and ethnic groups are more likely to live near hazardous waste sites and industrial facilities, and fenceline communities are typically subject to multiple forms of discrimination resulting in limited financial, political, and social capital to mitigate contaminant exposures7. Moreover, longitudinal analyses show that disproportionate hazard burdens faced by racially and economically marginalized groups are largely due to discriminatory land-use, permitting, and facility siting decisions8,9,10,11. Racial residential segregation and the inequitable distribution of stormwater infrastructure further contribute to racialized patterns of flood risk across U.S. cities12.Building upon a prior California analysis13, we conducted a nationwide equity assessment of flood risk at hazardous sites in the U.S. due to sea level rise (SLR). We derived probabilistic estimates of flood risk in 2050 and 2100 across an expanded range of legacy contamination sites and facilities that contain, handle, produce or emit hazardous substances. We then assessed the geographic distribution of at-risk sites with respect to multiple present-day measures of social marginalization, including race/ethnicity, poverty (household income below twice the federal poverty line), voter turnout, housing tenure, and linguistic isolation. Our objectives were to characterize inequities in residential proximity to hazardous sites at risk of future flooding due to sea-level rise and identify communities where additional resources are needed to prevent exposure to toxic substances and enhance climate resilience.ResultsHazardous sites at risk of floodingWe first assessed the annual probability of at least one flood exceeding the land elevation of over 47,646 coastal hazardous site locations compiled from one proprietary and four publicly available administrative data sources (Supplemental Table S1). We considered all sites within counties with land area below the 18 m elevation above current mean higher high water line across all coastal U.S. states and Puerto Rico. We defined sites as at risk if their projected annual probabilities exceeded 0.01 (i.e., they were threatened by a 1-in-100-year flood event) integrated across the full distribution of SLR projections using the law of total probability for one low (Reference Concentration Pathway [RCP] 4.5) and one high (RCP 8.5) greenhouse gas emissions scenario (see “Methods”).We found that over 11% of coastal sites in our analysis are at risk of SLR-related flooding by 2100 under the high emissions scenario (RCP 8.5) (Table 1). Figure 1 shows the distribution of at-risk sites by state or territory under RCP 8.5 in 2050 and 2100. Seven states (Louisiana, Florida, New Jersey, Texas, California, New York, and Massachusetts) account for nearly 80% of projected at-risk sites in 2100 (Fig. 1). Restricting greenhouse gas emissions to the low emissions scenario makes little difference in terms of the number of projected sites at risk in the near term (2050) but would reduce the number of at-risk sites from 5500 to 5138 (a reduction of 362 or 7% of sites) in the long term (2100) (Table 1). Oil and gas wells and industrial facilities that emit quantities of hazardous substances that require reporting to the U.S. Environmental Protection Agency’s Toxic Release Inventory (hereafter “TRI sites”) make up the largest proportion of sites we considered and sites at risk (Table 1). Under the high emissions scenario (RCP 8.5), over a fifth of coastal sewage treatment facilities, refineries and formerly used defense sites, roughly a third of power plants, and over 40% fossil fuel ports and terminals are projected to be at risk by 2100 (Table 1).Fig. 1: Number of sites at risk of flooding due to sea level rise in (left) 2050 and (right) 2100 under a high emissions scenario (RCP 8.5) by state and type.States are shaded by the total number of at-risk sites, with darker colors representing a higher number of sites at risk (maps). The number of sites at risk in each state is broken down by type, with each facility type represented by a unique color (bar chart).Full size imageTable 1 Number and type of hazardous sites at risk of sea level rise-related flooding by greenhouse gas emission scenario and year across the coastal U.SFull size tableAffected communitiesWe next considered the distribution of at-risk sites with respect to community demographics and indicators of social marginalization derived from three secondary datasets: the American Community Survey, a proprietary data source on recent voter turnout, and the federal Climate and Economic Justice Screening Tool14. We utilized census block groups as the geographic unit of analysis (hereafter “neighborhoods”) and considered block groups with at least one at-risk site located within 1 km of a populated area as being potentially affected (see “Methods”). Given the prominence of racial discrimination as a means of establishing and maintaining social inequality in the U.S.15, we considered measures of racial and ethnic makeup, as well as indicators of socioeconomic status, civic engagement (voter turnout), and vulnerability that relate to communities’ ability to anticipate, mitigate, and cope with flooding, such as age, linguistic isolation (% of households where no one 14 years or older speaks English “very well”), and vehicle ownership.Table 2 summarizes the population characteristics of neighborhoods near versus far from hazardous sites at risk of flooding due to SLR in 2100 under a high emissions scenario. Figure 2 shows the increase in the likelihood of an at-risk site within 1 km per one standard deviation increase in each demographic and social vulnerability measure, which we estimated using logistic regression models controlling for population density and county to minimize bias related to the higher concentration of people of color and renters in urban areas and demographic variation across U.S. regions.Fig. 2: Odds ratios and 95% confidence intervals for the association between the presence of socially marginalized groups and the likelihood of an at-risk site within 1 km under RCP 8.5, 2100 among coastal neighborhoods (N = 51,957 block groups).Black circles are adjusted odds ratios from models that considered one population characteristic at a time and controlled for population density and county fixed effects. Error bars indicate 95% confidence intervals and were calculated using robust standard errors. The dashed line indicates no association. Disadvantaged status (as defined by the federal Climate and Economic Justice Screening Tool [CEJST]) is a binary variable; all other variables are continuous and were scaled by unit standard deviation to facilitate comparisons between effect estimates.Full size imageTable 2 Characteristics of coastal block groups (n = 51,772) with and without at-risk sites within 1 km of populated areas in 2100 under RCP 8.5 across the U.SFull size tableCompared to other coastal neighborhoods, neighborhoods with one or more at-risk site nearby have lower voter turnout, proportions of residents identifying as Asian/Pacific Islander, and individuals under the age of 18, and higher present-day proportions of renters, households living in poverty, residents identifying as Hispanic and Black, linguistically isolated households, households without a vehicle, single-parent households, and individuals over the age of 65 (Table 2). In the multivariable regression models, all these bivariate associations remained statistically significant with the exception of the proportion of Black and Asian/Pacific Islander residents (Fig. 2). Neighborhoods designated as disadvantaged by the federal Climate and Economic Justice Screening Tool, a nationwide composite assessment of cumulative impact associated with multiple measures of social vulnerability (e.g., poverty) and the presence of climatic and environmental hazardous, had a 50% higher odds of having an at-risk site within 1 km, compared to other coastal, non-disadvantaged neighborhoods (odds ratio [OR] and 95% confidence interval [CI] = 1.50 [1.23, 1.83], Fig. 2). A one standard-deviation increase in the proportion of residents over age 65, linguistically isolated households, residents identifying as Hispanic, households in poverty, households without a vehicle, non-voters, and renters was associated with 15–41% higher likelihood of an at-risk site within 1 km (ORs and 95% CIs shown in Fig. 2). Associations were similar when we considered the presence of at-risk sites within 3 km instead of 1 km of neighborhoods (Supplementary Table S2).Among neighborhoods within 1 km of an at-risk site, social marginalization was also associated with an increase in the number of at-risk sites nearby and the severity of flood risk across those sites (Supplementary Fig. S1). Here we quantify flood risk severity by estimating the neighborhood expected annual exposure (EAE), calculated by summing the annual probabilities of at least one flood occurring at all hazardous sites within 1 km of populated portions of census block groups. This expected value reflects the total number of sites likely to be exposed to flooding in a given year—either 2050 or 2100 (see “Methods”). Among neighborhoods with an at-risk site within 1 km, a one standard deviation increase in the proportion of Hispanic residents, households in poverty, households without a vehicle, non-voters, and renters was associated with a 7–13% higher number of at-risk sites in 2100 under RCP 8.5 and a 0.10–0.21 unit increase in EAE (see Incidence Rate Ratios, mean differences and corresponding 95% CIs in Supplemental Fig S1). These estimates again control for population density and county to minimize bias.Most inequitably distributed sitesFigure 3 presents concentration indices and 95% confidence intervals summarizing the degree of inequality in the distribution of at-risk sites with respect to demographic and social marginalization indicators. We utilized concentration indices to identify the categories of facilities that were the most inequitably distributed for particular populations. Similar to the Gini coefficient commonly used to characterize income inequality, a concentration index ranges from −1 to 1, with negative values (in orange) indicating that the burden of at-risk sites is disproportionately higher for more marginalized groups and positive values (in blue) indicating that the burden is disproportionately lower for marginalized groups. Values are shaded white when confidence intervals include the null, indicating no significant evidence of a disparity. Concentration curves corresponding to the indices given in Fig. 3 are included in Fig. 4. These display the distribution of at-risk sites with respect to demographic and social marginalization measures, with the area between the curve and diagonal line of equality being equivalent to the concentration index (e.g. between −1 and 1). To increase the legibility of Fig. 4, for each demographic and social marginalization measure we display only the five site categories with the strongest concentration indices, while the full set of concentration indices is shown in Fig. 3.Fig. 3: Concentration indices and 95% confidence intervals for the cumulative distribution of at-risk facilities with respect to selected demographic and social marginalization measures under RCP 8.5, 2100.Negative values (in orange) indicate a disproportionately higher burden of at-risk sites for marginalized groups, while positive values (in blue) indicate that the burden is disproportionately lower for these groups. White values indicate a lack of statistical significance at P > 0.05. No adjustments were made for multiple comparisons.Full size imageFig. 4: Cumulative distribution of the number of at-risk sites with respect to selected demographic and social marginalization measures under RCP 8.5, 2100.The X-axis gives the cumulative share of block groups in descending order of each of the demographic and social marginalization variables. Curves above the equality line indicate a disproportionately higher burden of at-risk sites for marginalized groups, while a curve below the equality line indicate that the burden is disproportionately lower for these groups. For example, the top left panel shows that the 50% of low-lying block groups with the highest proportion of renters (indicated by the x-axis value of 0.5) host roughly 80% of at-risk fossil fuel ports and terminals and industrial facilities (indicated by y-axis values of 0.8), whereas if these sites were equitably distributed, the y-axis value would be close to 0.5 and the curves would fall closer to the bolded diagonal line of equality. For legibility, the top 5 facility categories with the strongest concentration indices for each measure are shown.Full size imageAt-risk power plants, industrial TRI sites, clean-up sites, and fossil fuel ports and terminals disproportionately burdened neighborhoods with higher proportions of renters, non-voters, households without a vehicle, households living in poverty, and linguistically isolated households, as indicated by negative concentration index values in Fig. 3 and curves above the line of equality in Fig. 4. In contrast, at-risk concentrated animal feeding operations and active oil and gas wells more often did not disproportionately burden marginalized groups, as indicated by mostly positive concentration index values in Fig. 3 and curves below the line of equality for many panels in Fig. 4, although there were exceptions. At-risk refineries disproportionately burden neighborhoods with higher proportions of non-voters, households in poverty, and Black residents, and at-risk TRI facilities disproportionately burden neighborhoods with higher proportions of Black, Hispanic and Asian/Pacific Islander residents. In contrast, neighborhoods with a higher proportion of Native American residents are projected to be disproportionately burdened by at-risk active oil and gas wells, hazardous waste sites, landfills, and formerly used defense sites (Figs. 3,  4).Conclusions about which at-risk site types are inequitably distributed are largely but not entirely consistent across different metrics of flood risk (number of at-risk sites, which is presented in Fig. 3 vs. EAE across sites which is presented in Supplementary Fig S2). For example, neighborhoods with higher proportions of renters, linguistically isolated households, and households without a vehicle were not burdened by a disproportionate share of at-risk refineries (Fig. 3), but when assessing EAE, they were disproportionately burdened (Supplementary Fig S2 and S3). Similarly, Hispanic and Asian/Pacific Islanders were not burdened by a disproportionate share of at-risk refineries (Fig. 3), but are disproportionately burdened when considering EAE (Supplementary Fig S2 and S3). This may be because although the number of at-risk refineries tends to be higher near neighborhoods with smaller proportions of these residents, the severity of projected flooding at those refineries is higher than it is near other neighborhoods.DiscussionWe present a national assessment of projected SLR-related flooding threats to multiple categories of coastal sites and facilities that contain, use or produce hazardous materials. Our results show that of the more than 47,600 coastal facilities in the U.S. included in our analysis, over 11% (5500 facilities) are projected to be at risk of a 1-in-100-year or more frequent flood event by the end of the 21st century (2100) under a high (RCP 8.5) greenhouse gas emission scenario. A handful of states, including Louisiana, Florida, New Jersey, Texas, California, New York, and Massachusetts account for nearly 80% of projected at-risk sites.Facilities at risk include 22% of coastal sewage treatment facilities, 24% of refineries, 44% of fossil fuel ports and terminals, 12% of industrial facilities, 21% of formerly used defense sites and 30% of fossil fuel and nuclear power plants. A prior study estimated the number of wastewater treatment plants and service populations across the U.S. that could be exposed to SLR scenarios from 1 to 6 ft, with projections ranging from 60 impacted treatment plants serving 4 million people to 394 plants serving over 31 million people16. That analysis did not incorporate elevated water levels due to tides, waves, and storm surge, which likely explains why we projected a larger number of sewage treatment facilities to be at risk of SLR-related flooding in the RCP 8.5 scenario. Another prior assessment of how unmitigated greenhouse gas emissions could affect U.S. power-generating capacity in 2100 among power plants in coastal areas estimated a similar number of power plants at risk as we did in our analysis. That study additionally considered the generation capacity of at-risk power plants. The authors found significant variation across states with exposed power capacities relative to current generation capacities being highest in Delaware, New Jersey and Florida (80%, 63% and 43%, respectively)17.Our analysis shows that industrial facilities that are part of the Toxic Release Inventory make up nearly a third (34%) of the total sites at risk of SLR-related flooding (N = 1870), second to fossil fuel infrastructure (41%), including refineries, fossil fuel ports and terminals and active oil and gas wells. Because we did not include pipelines in our analysis, our projections of the extent to which the nation’s fossil fuel infrastructure may threaten coastal communities due to SLR-related flooding and associated contaminant releases are likely an underestimate. Indeed, extreme weather events such as Hurricanes Katrina, Rita and Harvey, while very different from the slower moving, incremental flooding related to SLR, have dramatically revealed the vulnerability of industrial facilities and oil and gas infrastructure. Flooding following these hurricanes led to oil and chemical spills, pipeline ruptures, as well as excess air pollutant emissions due to incidental releases as well as intentional shutdowns, flaring, and subsequent restarting of operations at petrochemical facilities18,19,20,21,22,23. Our prior equity analysis of contaminant releases related to Hurricanes Harvey, Rita and Ike found that these natural-technological (natech) disasters disproportionately impacted Hispanic, renter, low-income, and rural populations5. Similarly, results in this study show significant inequities in projected SLR flooding threats to potentially hazardous facilities; communities defined as disadvantaged by the federal Climate and Economic Justice Screening Tool (CEJST) have a 50% higher odds of having an at-risk site within 1 km, and a one standard deviation increase in the proportion of linguistically isolated households, neighborhood residents identifying as Hispanic, households in poverty, households without a vehicle, non-voters, and renters is associated with 19–41% higher likelihood of an at-risk site.Our findings align with prior equity studies of current and projected distributional burdens of flood risk among diverse populations in the U.S. A national study using Federal Emergency Management Agency maps from 2001–2019 in urban areas along with National Land Cover Data and county-level Census data found that 100-year flood zones, particularly in coastal counties, are often occupied by a higher proportion of disadvantaged populations24. Another study of coastal and inland areas estimated an average increase of 26.4% (24.1–29.1%) in climate change related flooding by 2050 under an RCP4.5 scenario, with the future increase in flooding risk concentrated on the Atlantic and Gulf coasts and disproportionately affecting Black communities25; although this study examined flooding and economic losses related to residential and non-residential properties, it did not consider risks to potentially hazardous sites. Other studies have examined flooding threats to active and legacy sites containing hazardous material. For example, a report found low-income communities were disproportionately represented among the populations living in proximity to clean-up sites (listed or candidate sites for the Superfund program) at risk of coastal flooding under low, medium, and high SLR scenarios in the East and Gulf Coasts26. A follow-up study identified coastal land below 10 m of elevation as potentially exposed to rising groundwater and identified 326 Superfund sites in these coastal areas that could experience mobilization of toxic compounds from contaminated soil due to groundwater inundation driven by SLR; results also showed that socially marginalized groups in several states would be disproportionately affected by this groundwater rise scenario27. Another analysis of former hazardous manufacturing facilities in six U.S. cities identified more than 6000 relic industrial sites with elevated flood risk over the next 30 years (2050), with socially vulnerable groups, including people of color and low income, disproportionately likely to live in these areas28. Studies outside of the U.S., for example in coastal regions in India, Copenhagen, Vietnam and Italy have investigated the risks posed by climate change-driven SLR and storm surge on infrastructure and vulnerable sites29,30,31,32, but none to our knowledge have evaluated these risks using an environmental justice framework.Strengths of our study include the use of tax parcel data to characterize the extent of facility boundaries, a probabilistic approach to estimating SLR-related flood risk, and the application of dasymetric mapping techniques to estimate populations and community demographics near at-risk sites. Limitations of our analysis include the fact that our flood models assume that the frequency and magnitude of flood events will remain static over the next century. However, studies indicate that tropical cyclone activity is likely to intensify due to the acceleration of climate change33,34,35, which would result in more damaging impacts to coastal communities36. Additionally, our estimations of annual probabilities of flood level exceedance, based on a modified “bathtub” approach, do not consider scenarios of groundwater intrusion and upwelling or nonlinear interactions between extreme flood events and local topography. These dynamics could cause increased flood levels at inland locations, especially where marshlands shrink, and land becomes more developed37. Our analysis also does not account for floodwater level attenuation particularly in areas where land is wide and flat, which may overestimate exposure during extreme storm events38. Locational errors for hazardous sites may have also led to over- or under-estimates of the number of at-risk sites, and data limitations precluded inclusion of other facility types, including underground storage tanks, brownfields, and non-National Priority List Superfund sites that could experience contaminant releases due to SLR-related flooding. Inaccuracies in the delineation of coastline boundaries may have resulted in the inclusion of offshore drilling sites and the overestimation of flood risk. Finally, we did not account for future flood risk mitigation efforts or population and demographic shifts, given the high degree of uncertainty in predicting these scenarios. Therefore, future actions to mitigate flood risk near hazardous sites, gentrification, displacement, migration, and other factors could change the associations we observed between demographics, measures of social marginalization, and proximity to at-risk sites.Our analysis highlights the disproportionate burden of projected SLR-related flooding threats to hazardous sites on marginalized racial and socioeconomic groups and elevates the importance of centering environmental justice in future climate change adaptation and land-use planning strategies to protect vulnerable coastal communities from natech disasters. Given that nearly 80% of projected at-risk sites are in seven states, future in-depth work can target these areas and more precisely characterize the potential hazards posed by these facilities to nearby communities with the goal of mitigating and preventing future harmful exposures and health risks. With over 30% of nuclear and fossil fuel power plants, 23% of refineries, and 44% of fossil fuel ports and terminals in coastal areas projected to be at risk, federal reporting requirements for these facilities could be expanded to include the forecasting of SLR-related flooding threats and preventive plans for mitigation, including future relocation, to avoid catastrophic contamination. Critical to these efforts will be ensuring that federal and state agencies provide publicly available, accessible, and continually updated data on projections of SLR-related flooding threats to hazardous sites for diverse end-users, in particular at-risk communities, planners, regulatory agencies, scientists, and decision-makers39. Future research focusing on a smaller subset of facilities and more localized regions or municipalities could further elucidate and potentially untangle the extent to which place-based trends in industrial, economic, labor market, and housing development trajectories, demographic churning, changes in land-use decision-making as well as other shifting structural factors account for the origins and persistence of inequities in exposures to at-risk sites that disproportionately affect marginalized populations.Finally, many other climate-related phenomena, such as groundwater rise, wildfires, landslides, major storms, and extreme heat, also threaten clean-up sites and active facilities that use and store hazardous materials27,40,41. To achieve a fuller picture, information on these threats should be integrated with projected SLR flood risk data. Risks may be reduced through enhanced regulatory requirements (1) for at-risk facilities to mitigate and prevent contamination threats and (2) for more robust assessment of clean-up sites to inform abatement activities and decisions about future land reuse. Action-oriented partnerships between communities living near at-risk sites and government agencies at local, state, and federal levels may increase the chances for success of these strategies by marshalling much-needed resources aimed at preventing contamination from acute natech disasters and slower-moving threats, including SLR-related flooding.MethodsOur analytic approach entailed four steps: 1) the identification of coastal hazardous site locations and the cleaning of associated descriptive data; 2) the estimation of future flood risk due to sea level rise at each site location; 3) the compilation of measures of demographics and social marginalization; and 4) a neighborhood-level analysis of the relationship between these measures and residential proximity to at-risk sites. We co-developed these methods with an advisory committee comprised of staff members from environmental justice and public health organizations with whom we collectively decided on greenhouse gas emissions scenarios, timeframes (2050 and 2100), flood risk metrics, categorization of sites, and the demographic and social vulnerability metrics to include13.Hazardous sitesThe spatial extent of our analysis was U.S counties and county equivalents with any land area below 18 meters elevation above current mean higher high water line across all coastal U.S. states and Puerto Rico (see Supplementary Fig S4). Areas farther inland are at no conceivable risk of flooding due to sea level rise this century and were therefore excluded. We scaled up an approach for a prior analysis of California13 to compile a national dataset of active industrial facilities and other potentially hazardous sites. To achieve this, we sourced data from the U.S. Environmental Protection Agency’s (EPA) Facility Registry Service (FRS)42, the U.S. Energy Information Administration’s (EIA) Energy Atlas43 (petroleum refineries and terminals), the U.S. Army Corp of Engineers’ (USACE) Waterborne Commerce Statistics Center44 (petroleum ports) and Formerly Used Defense Sites database45, and a proprietary dataset of active oil and gas production and stimulation wells from EnverusTM 46. For the FRS, we chose to exclude remediated and closed facilities and facilities with inaccurate or imprecise locational information (e.g. latitude and longitude values derived from zip codes only or with inaccuracy >50 m). This included sites with environmental interest “end dates” indicating they would no longer be regulated after 2020 or where records indicated contamination had been addressed or the site was permanently closed. We retained inactive facilities and facilities with expired permits because residual hazardous materials may remain at these sites. We organized the remaining sites into one of seven categories using (1) environmental permits or regulatory programs; (2) the North American Industry Classification System (NAICS) code; and/or (3) keyword filters (see Cushing et al. 13 for further detail). We made sure that each category was mutually exclusive and without duplicate entries, as sites can have more than one environmental permit and/or NAICS code and appear in more than one database. We manually removed FRS entries for refineries using refinery names and coordinates from the EIA Energy Atlas dataset.For oil and gas wells, we utilized latitude and longitude point locations to represent sites due to the small size of well pads compared to other site categories. We identified offshore oil and gas wells as those that were beyond the boundaries of 2010 Census block groups, and excluded them from the analysis to focus on wells located on land. Block group boundaries from the National Historical Geographic Information System do not include coastal water areas and terminate at the coastline. All other site types were represented as polygons in our analysis to better approximate a site’s extent. For FRS, EIA, and USACE petroleum port sites, we used the Google API47 to (re)geo-code addresses, then joined the resulting coordinates to tax parcels obtained from Loveland Technologies (now Regrid)48 to approximate their spatial boundaries. Wherever a site’s geocoded location overlapped with a tax parcel, we used that parcel to approximate that site’s spatial extent for the purposes of projecting flood risk. Around 79% of our geocoded site locations fell within tax parcel boundaries. For sites that did not intersect tax parcels, we approximated boundaries by drawing a buffer equal to the median parcel area of intersected parcels for each site category (See Supplementary Material Table S1 for median areas applied for each category). For formerly used defense sites (FUDS), spatial data were available in both point and polygon format. Not all sites had a point or polygon, while some sites had both. To ensure we included all FUDS in our dataset, we (i) first included all facilities with polygon data, then (ii) identified facilities with point locations falling outside of these provided polygon boundaries, and (iii) drew a buffer around these point facilities using a circular radius that would result in the median area observed among facilities in (i). Wherever a polygon boundary overlapped with a buffered point, we clipped the latter based on the physical extent of the former (n = 25). For overlaps between two polygon boundaries, we used an overlap ratio (calculated as the area of the overlap divided by the area of the smaller polygon) to determine whether to split the overlapping area evenly between two facilities with minimal overlap (≤45%, n = 33), or to merge substantially overlapping facilities together into one (>45% overlap, n = 42). We resolved 13 complex cases involving three or more overlapping facilities manually on a case-by-case basis. The result of this process was a dataset of FUDS boundaries derived from original point locations or polygon boundary extents that contained no spatial overlaps between sites. For all sites, we then clipped parcels and circular buffer areas at the coast if they extended past the mean high tide line.As a final cleaning step, we flagged and subsequently consolidated duplicate sites (n = 656) if they met three criteria: they were assigned to the same category, had identical geo-coded coordinates, and were associated with the same or similar addresses (we quantified similarity using a fuzzy text match). We retained facilities with identical coordinates and similar addresses if they had been assigned to different categories (n = 232). We dropped facilities with matching coordinates but dissimilar addresses (n = 30) if their geocoded coordinates appeared inaccurate or implausible via manual visual inspection (e.g., located in the middle of a forest far away from any established roads).Flood risk projectionsWe used the same approach to assess flood risk at individual site locations as detailed in Cushing et al. (see Supplementary Fig S5)13,49,50. In brief, we considered probabilistic sea level rise projections51 for two greenhouse gas emissions scenarios (Reference Concentration Pathway [RCP] 4.5 and 8.5) and 2 years (2050, 2100)51. For each site, year, and emissions scenario, we estimated the total annual probability of at least one flood event exceeding, in height, the 25th percentile of land elevation for a given site’s parcel or buffer boundary. Projections account for vertical land movement and coastal flood height return level curves using methods from Tebaldi et al. and updated tide station data from across the United States52. We derived elevation profiles from NOAA’s Coastal Topographic Lidar digital elevation model53, and estimated the annual flooding probabilities using Equation (1) from Buchanan et al.50 We considered sites to be at risk if their projected annual probabilities exceeded 0.01 (i.e., threatened by a 1-in-100 year flood event). We also summed these probabilities across sites to derive a total expected annual exposure (EAE) across all at-risk sites within a given distance of neighborhood (block group) boundaries.Neighborhood demographics and social marginalizationWe used 2010 U.S. Census block group boundaries as our definition of geographic neighborhoods. Census block groups are generally contiguous geographic areas that contain between 600 and 3000 people and are the smallest unit for which the U.S. Census Bureau reports a full range of demographic statistics. We used the U.S. Census American Community Survey’s (ACS) 2015–2019 five-year estimates54 to approximate demographic characteristics at the block group level: age (% under 18 and % 65 and older), race/ethnicity (% people of color, defined as the inverse of % non-Hispanic White and disaggregated into % Hispanic, and % non-Hispanic [NH] Black, NH Asian or Pacific Islander, NH Native American, and NH other including multiracial), poverty (% below twice the federal poverty line), housing tenure (% renter-occupied units), vehicle ownership (% of households without a vehicle), family structure (% single parent-headed households), linguistic isolation (% of households where no one 14 years or older speaks English “very well”). We used voter turnout data from the 2016 and 2020 general elections from Catalist’s National Database to approximate civic engagement (% of registered voters that did not vote averaged across the two elections). Finally, we used the federal Climate and Economic Justice Screening Tool (CEJST) that identifies disadvantaged communities in all 50 states, the District of Columbia, and U.S. Territories55. Developed as part of the Justice40 Initiative, CEJST was used by federal agencies to identify disadvantaged communities facing disproportionate climate and environmental burdens as well as economic marginalization. CEJST identifies disadvantaged communities through eight categories of vulnerability metrics related to climate change, energy, health, housing, legacy pollution, transportation, water and wastewater, and workforce development. Census tracts are identified as disadvantaged if they meet 90th percentile thresholds for indicators within any of the eight categories and are at or above the 65th percentile for low-income.Statistical analysisWe began by identifying and including only counties in our study area with at least one site at risk under RCP 8.5 by 2100. We then further restricted the geographic extent of our analysis to “coastal” block groups in these counties within 3-km Euclidean distance of the 10-m elevation line above mean higher high water line. Our primary outcome of interest was the presence of at least one at-risk site within 1 km. We considered block groups to have this outcome if they contained populated areas within a kilometer of at least one at-risk site (see Supplementary Fig S4). Because block groups can be quite large in rural areas, we utilized gridded population estimates at a 30 × 30 m resolution56 to define populated portions of block groups with the exception of Alaska, Hawaii and Puerto Rico for which these estimates were not available and where we therefore relied on block group boundaries alone. We secondarily considered (1) the total number of at-risk sites within 1 km, and (2) the sum of annual flood event probabilities (total “expected annual exposure”, EAE) across all at-risk sites within 1 km. We conducted sensitivity analyses considering alternate versions of these outcomes using a 3 km rather than 1 km buffer distance.We examined descriptive statistics and correlation coefficients between our demographic measures and indicators of social marginalization and compared the distribution of neighborhood characteristics between exposed and unexposed block groups using Mann-Whitney U test because variables were not normally distributed. We then ran multivariable regression models for each outcome variable and vulnerability indicator pair, with block-group population density (people per square kilometer), and county fixed effects as additional independent variables. We did not include multiple demographic or social marginalization indicators in the same model due to multicollinearity. We chose to include population density as a potential confounder due to known associations between race/ethnicity, population density, and proximity to industrial facilities57,58. We included county fixed effects to control for regional demographic differences and in effect compare block groups with and without at-risk sites within the same county. We scaled continuous variables by unit standard deviation (SD), using the mean and SD from all block groups in our universe to allow for easier comparisons between effect estimates. In our primary analysis, we used a logistic model to estimate the odds of proximity to an at-risk site (yes/no variable). Restricting to exposed block groups, we used a negative binomial model to estimate associations with the number of sites nearby (count variable) and a linear model to estimate associations with EAE (continuous variable). We used county-clustered robust standard errors to control for the spatial autocorrelation.Finally, we used concentration curves to visualize the cumulative distribution of the number of exposed facilities and EAE with respect to each indicator of social marginalization. We also derived the concentration index (C) equal to the area beneath the curve and line of equality in our concentration plots to quantify the cumulative distributions. C ranges between −1 and 1, with negative values indicating that block groups with higher proportions of residents from socially marginalized groups have a greater number of at-risk facilities and EAE, and positive values indicating they have a smaller burden of at-risk sites and EAE. A C value with a confidence interval that includes the null value of 0 indicates that the number of exposed facilities and EAE are similar between more and less marginalized populations. We calculated C using all coastal block groups, and we calculated separate indices for each facility category focusing on year 2100 under RCP 8.5. Concentration curves and indices were computed using R (version 4.5.0). All other statistical analyses were conducted using Python (version 3.13.7).Reporting summaryFurther information on research design is available in the Nature Portfolio Reporting Summary linked to this article.

    Data availability

    The datasets generated during the current study are available from the Toxic Tides maps in Climate Central’s Coastal Risk Screening Tool (flood risk projections, https://coastal.climatecentral.org/), GitHub (analytic dataset and code, https://github.com/yangju-90/toxic_tides_us), and Zenodo (analytic dataset and code, https://doi.org/10.5281/zenodo.16925499).
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    Download referencesAcknowledgementsThis project has been funded wholly or in part by the United States Environmental Protection Agency (EPA) under assistance agreement 84003901 to L.J.C. The contents of this document do not necessarily reflect the views and policies of the EPA, nor does the EPA endorse trade names or recommend the use of commercial products mentioned in this document. We thank the Toxic Tides Advisory Council—comprised of community leaders from the Asian Pacific Environmental Network, Central Coast Alliance for a Sustainable Economy, Physicians for Social Responsibility Los Angeles, Public Health Institute, and WE ACT for Environmental Justice—for advising on the methods.Author informationAuthors and AffiliationsDepartment of Environmental Health Sciences, University of California, Los Angeles, CA, USALara J. Cushing & Alique BerberianDepartment of Land Resources and Tourism, School of Geography and Ocean Science, Nanjing University, Nanjing, ChinaYang JuEnergy and Resources Group, University of California, Berkeley, CA, USASeigi Karasaki & Nicholas DepskyClimate Central, Princeton, NJ, USAScott Kulp & Benjamin StraussSchool of Public Health, University of California, Berkeley, CA, USAJessie Jaeger & Rachel Morello-FroschDepartment of Environmental Science, Policy and Management, University of California, Berkeley, CA, USARachel Morello-FroschAuthorsLara J. CushingView author publicationsSearch author on:PubMed Google ScholarYang JuView author publicationsSearch author on:PubMed Google ScholarSeigi KarasakiView author publicationsSearch author on:PubMed Google ScholarScott KulpView author publicationsSearch author on:PubMed Google ScholarNicholas DepskyView author publicationsSearch author on:PubMed Google ScholarAlique BerberianView author publicationsSearch author on:PubMed Google ScholarJessie JaegerView author publicationsSearch author on:PubMed Google ScholarBenjamin StraussView author publicationsSearch author on:PubMed Google ScholarRachel Morello-FroschView author publicationsSearch author on:PubMed Google ScholarContributionsL.J.C. contributed to manuscript writing and jointly conceived of the project, acquired the funding, and supervised the work. Y.J. conducted the statistical analysis and reviewed and edited the manuscript. S Karasaki contributed to data curation and manuscript writing, and prepared figures and tables. S.Kulp conducted the flood risk projections and reviewed and edited the manuscript. N.D. and J.J. contributed to data curation. A.B. contributed to data curation, preparation of figures and tables, and edited the manuscript. B.S. jointly conceived of the project and acquired the funding, and reviewed and edited the manuscript. R.M.F. contributed to manuscript writing and jointly conceived of the project, acquired the funding, and supervised the work.Corresponding authorsCorrespondence to
    Lara J. Cushing or Rachel Morello-Frosch.Ethics declarations

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    Reprints and permissionsAbout this articleCite this articleCushing, L.J., Ju, Y., Karasaki, S. et al. Sea level rise and flooding of hazardous sites in marginalized communities across the United States.
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    Mandated on-site wastewater reuse in San Francisco: the role of distributive fairness for policy acceptance

    AbstractWith increasing water scarcity worldwide, policies regulating wastewater reuse are becoming increasingly important. In San Francisco, on-site wastewater treatment and reuse is mandatory for large residential buildings, while other buildings continue using centralised systems without reuse. This disparity may affect perceived fairness and policy acceptance. In an online survey (N = 176), policy acceptance, perceived fairness, and perceptions of a range of policy implications were assessed for five societal groups and one entity: residents and owners of buildings with mandated on-site systems, San Francisco’s population, low-income residents, future generations, and the environment. Regression analyses showed that both positive and negative policy implications explained perceived fairness. Policy acceptance was explained by perceived fairness for future generations, San Francisco’s population, and building owners, but not other groups or entities. Results suggest that collective fairness considerations and impacts on most-affected groups are key to policy acceptance, indicating policymakers should consider implications across different societal groups when designing water reuse policies.

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    IntroductionWater scarcity poses a critical challenge in many urban settings globally. With increasing population growth and intensified climate change, the strain on water resources becomes increasingly apparent1,2, necessitating innovative solutions to manage water systems effectively. One such solution gaining traction is on-site wastewater treatment for non-potable reuse3,4. For this approach, all or part of the wastewater generated within a single building or a small cluster of buildings is treated on-site. The water can then be reused on the buildings’ premises, for example, for irrigation, clothes washing, or toilet flushing. Such on-site wastewater treatment and reuse (hereinafter referred to as on-site reuse) can reduce the demand on a city’s drinking water resources5.To encourage the uptake of systems for on-site wastewater treatment and reuse (on-site systems) in water-stressed regions, some local governments, among them San Francisco in the United States, Bengaluru in India, and Sant Cugat del Vallès in Spain, have enacted policies mandating the installation and use of on-site systems in specific (mostly large) buildings6,7,8. Yet, for an effective and lasting implementation of such policies and environmental policies in general, they need to be seen as legitimate and accepted by the public9. A lack of public acceptance (i.e. a positive valuation of a policy by the general public10,11,12) can initiate public opposition that may eventually lead to the failure of political measures. In fact, the implementation of several wastewater reuse projects in the United States and Australia failed as a result of local public protests13,14,15,16,17 (these projects aimed for treatment in centralised, not on-site treatment plants).While a lack of public acceptance of environmental policies may have various causes (for overviews, see refs. 9,18), one core reason has been shown to be a perceived lack of the policies’ distributive fairness (i.e. the fairness of the distribution of costs, risks, and benefits within or between different groups or entities of society19,20, also often referred to as ‘equity’21,22)9,23. A meta-analysis on determinants of policy acceptance and support of climate-change related policies revealed that perceived distributive fairness was the strongest determinant, stronger than, for example, perceived personal fairness, perceived policy effectiveness, or trust in political and implementing institutions9. Several experimental studies could further show that a perceived lack of distributive fairness consistently was the strongest barrier of public support for a range of environmental policies, stronger than a perceived lack of procedural fairness or extensive costs23. Also in the specific case of wastewater reuse, public protests have been associated to distributive fairness. Public opposition against centralised wastewater reuse schemes in San Diego has in part been attributed to a perceived injustice, particularly against low-income groups13,24. Further, a study by Nancarrow et al.25 showed that a planned scheme for centralised reuse in Australia was less acceptable the more it was perceived as unfair to the local population as well as to a variety of groups of society. These fairness ratings were, however, averaged, forming an overarching measure of perceived fairness for societal groups.For policies mandating on-site (in contrast to centralised) wastewater reuse, distributive fairness may be particularly relevant26. Such policies usually extend to large buildings only to increase cost-effectiveness and environmental impact, which threatens an equal distribution of the policies’ costs, risks, and benefits among society. Specifically, the costs and risks primarily fall upon the owners or residents of buildings subject to these regulations, while it is likely that the whole local population benefits from the conservation and diversification of public water resources. Alternatively, depending on the quality of the city’s water supply, it is also possible that residents living in buildings with on-site reuse have a better water quality or increased quantity of water than all other citizens. In both cases, the resulting unequal distribution of costs, risks, and benefits sets the fair provision of water services within a society at risk. Specifically, the principle of equality may be at risk as the policy will not equally impact all individuals27,28. In fact, a recent study with residents of Bengaluru, India, has shown that if people covered by the local policy mandating on-site reuse are perceived to have a worse outcome compared with people not covered, the policy is perceived as less fair and acceptable29.While research has shown that people aim for distributive fairness of policies, the policy does not necessarily need to have an equal outcome (i.e. an equal distribution of related costs, risks, and benefits) for all individuals in order to be perceived as fair by the public. People give more weight to the outcome for some groups or entities of society than for others. For example, Schuitema et al.30 found that perceived fairness and acceptance of a policy mandating transport pricing was more strongly predicted by considerations of the policy’s outcome for the environment and future generations than by considerations of people’s personal outcome.Other studies have shown that special attention is also paid to outcomes for members of society that are marginalised and in need. This is in line with the fairness principle of ‘need’27,31, denoting that an allocation of outcomes based on people’s need is perceived as fair. For instance, the above-mentioned study on transport pricing30 showed that the policy measures were perceived as less fair and less acceptable when low-income groups were burdened. A review on carbon pricing policies32 and a recent study33 reported lower acceptance the more low-income groups were burdened, and a study by Andor et al.34 indicated that acceptance of a levy on renewable energies for electricity costumers was higher when low-income groups were exempt from it. Further, a study by Pitkänen et al.35 indicated that also other groups of society that are in need are considered for fairness evaluations: Participants perceived a policy on personal carbon trading as unfair if it burdened people with high mobility needs, such as families with children or rural households.To our knowledge, only one study29 investigated which different groups or entities of society are considered for perceived fairness and acceptance of a policy mandating on-site wastewater reuse. This study, conducted in India, showed that among participants who are covered by the policy, perceived fairness and acceptance of the policy was predicted by considerations of the policy’s outcome for the environment and future generations but not by their personal outcome, which is in line with Schuitema et al.30. However, in contrast to the studies described above30,32,33,34,35, Kollmann et al.29 found that the policy’s outcomes for low-income groups or those vulnerable to water insecurity were not considered when assessing fairness and acceptability. Yet, the outcome for people covered by the policy explained fairness and, only among people not covered by the policy, also acceptance.It is unclear whether the findings by Kollmann et al.29 and their deviation from previous findings regarding groups in need indicate a general pattern for policies that mandate on-site wastewater reuse or that cover only part of the population or whether they are context-specific to India. There are arguments in favour of both interpretations. What could suggest a general pattern, is that in the context of policies mandating the use of a technology (compared with pricing policies) the outcome for low-income groups may be less relevant, as such policies involve not only monetary but also behavioural costs. Moreover, for policies covering only part of the population, only the outcome for those covered may be considered, while additional disparities caused for people in need may be considered less. What could, however, support a context-specific interpretation, is the fact that India is a lower middle-income country36, while the studies found that vulnerable groups of society being considered for fairness and acceptance ratings30,32,33,34,35 were all conducted in European high-income countries. Residents of a lower middle-income country might be less able to ‘afford’ considering the implications for people with low incomes than residents of high-income countries. In addition, also the cultural context in India may have had an influence. The Indian caste system, though abolished, still influences the perception of social inequalities, potentially making them more accepted37. As it is unclear why groups in need were not considered by participants in the study by Kollmann et al.29, it is important to investigate in a context that differs in many aspects with India if and to what extent perceived fairness for different groups of society, including those in need, can explain acceptance of a policy mandating on-site reuse for part of the population.In addition, it is equally relevant to better understand why a policy is perceived as (un)fair for the specific groups of society in the first place. Understanding which specific implications of a policy lead people to evaluate it as (un)fair for a certain group can help designing future policies in a way that they are perceived as fair and acceptable by the public. Previous research on environmental policies showed that they are considered fairer if they consist of pull mechanisms that incentivise or reward pro-environmental behaviour rather than of push mechanisms disincentivising certain behaviour or imposing costs30,38. Moreover, policies encouraging upward social mobility39 or redistributing their revenues to environmental projects40 are considered fairer. Yet, only a limited number of studies investigated which aspects of a policy design can explain its perceived fairness, and for the context of wastewater reuse, it has, to our knowledge, not been investigated at all. It is, for example, possible that financial implications of a policy are more or less important for fairness ratings than non-financial ones. Further, it is also possible that perceived benefits of a policy are considered more than perceived costs and risks. Kollmann et al.41 showed that perceived benefits of on-site systems are more important for the acceptance of on-site reuse than their costs and risks. It is possible that this pattern also emerges for the perceived fairness of a policy mandating on-site reuse.Although (on-site) wastewater reuse has become increasingly important around the world4,42 and with it policies regulating its implementation6, so far only one study29 has investigated whether fairness of such a policy explains policy acceptance and which groups of society are considered for fairness and acceptability assessments. In addition, it has not at all been investigated which specific implications of such a policy explain whether the policy is perceived as fair or unfair for different groups.Moreover, to our knowledge, neither with regard to such policies nor with regard to environmental policies in general have previous studies investigated whether and how strongly the perceived fairness for several groups of society explains policy acceptance. Previous studies have investigated whether the outcome of a policy (i.e. its costs, risks, and benefits) for different groups of society explain the policy’s perceived overall fairness, but without assessing individual fairness ratings for each group29,30,35. Studies that have included individual fairness ratings did so either only with regard to people themselves and for others in general, but not for more groups of society43 or created mean scores after the assessment resulting again in a measure of overall fairness25. While an overall fairness measure is relevant for the investigation of many research questions, it is too broad for analysing how the perceived fairness for different groups can individually explain policy acceptance and why a policy is rated as (un)fair for each group.In the present study, we therefore investigated (1) whether and to what extent the perceived fairness of the policy for five different societal groups and one entity (the environment) can explain acceptance of the policy, and (2) whether and to what extent different implications of the policy for each group and the environment explain perceived fairness of the policy for the specific group or entity. In additional exploratory analyses, we investigated whether perceived financial implications are more or less important for fairness ratings than non-financial ones and whether perceived benefits are more or less important than perceived costs and risks.These relationships were investigated in San Francisco where a policy is in place mandating the installation and use of on-site systems for certain building types and thus only for part of the population. The following five groups of society and one entity were investigated in the study: residents of buildings with mandated on-site systems, owners of buildings with mandated on-site systems, the city of San Francisco and its population, low-income residents of San Francisco, future generations living in San Francisco, and the local and regional environment.ResultsPolicy implications explaining perceived fairness for different societal groups and the environmentWe first assessed which policy implications explain perceived fairness of the policy for different societal groups or entities. All results of the regression analyses as well as the means and standard deviations are displayed in Tables 1–2. The bivariate correlations of all variables and the multicollinearity diagnostics for the regression analyses are presented in the Supplementary Information (Supplementary Tables 1–9).Table 1 Regression analyses explaining perceived fairness for different groups of societyFull size tableTable 2 Regression analysis explaining policy acceptance of on-site systemsFull size tableOf the eight policy implications for residents, four explained perceived fairness of the policy for this group of society. Two of these implications were on average perceived as negative by participants, namely that residents have to bear monetary costs related to the on-site systems and that residents might face reduced maintenance and water quality (compared with users of the centralised systems) as on-site systems are not managed by the city’s utility. The more these two implications were perceived as negative, the more the policy was perceived as unfair for residents. The other two policy implications explaining perceived fairness for residents were on average perceived as positive. These were that residents are likely to have a positive and sustainable image due to living in a ‘green’ building and that residents will have financial savings due to a lower freshwater consumption and potentially also due to a reduced fee to be paid to the utility for wastewater treatment. The policy was perceived as fairer for residents the more positively these two implications were viewed.Of the five policy implications for owners of buildings with mandated on-site systems, four explained perceived fairness of the policy for this group. Two of them were on average perceived as negative: The financial costs for installing on-site systems as well as the financial risk in case of technical issues that is carried by the building owners. The policy was perceived as more unfair for building owners the more these two implications were seen as negative for them. The other two implications explaining perceived fairness for building owners were on average perceived as positive: Financial savings for building owners due to a reduced fee to the utility for wastewater treatment and a positive and sustainable image of building owners for building or renting out units in a ‘green’ building. The more positive these two implications were perceived, the more fair was the policy perceived for building owners.Of the seven policy implications for the city of San Francisco and its population, three explained perceived fairness of the policy for this group of society. One of these implications was on average perceived as negative by participants, namely the possibility that the city’s centralised water and wastewater system becomes more expensive for users of the centralised system run by the utility, as residents of on-site system pay a reduced fee to the utility. The more participants perceived that this implication is negative for the city of San Francisco and its population, the more was the policy perceived as unfair for this group. The other two implications were on average perceived as positive by participants. These were that the policy leads to an increased wastewater treatment capacity of the city by taking load off the centralised system and that new, green jobs are created in San Francisco, as the policy increases the demand of on-site systems and thus builds a local, green economy. The more these implications were viewed as positive for the city of San Francisco and its population, the fairer was the policy perceived for this group.Of the four policy implications for low-income residents of San Francisco, three explained perceived fairness of the policy for this group of society. One of them was on average perceived as negative, namely that most low-income residents cannot profit from the monetary and non-monetary benefits of on-site reuse as most low-income residents will not live in buildings with on-site systems (due to the exemption of low-income housing developments from the policy). The more this implication was perceived as negative, the more was the policy perceived as unfair for low-income residents of San Francisco. The two other implications explaining perceived fairness for this group were on average perceived as positive by participants. These were that low-income housing continues to be built in San Francisco (due to the exemption, these projects are less expensive than conventional projects) and the implication that rents for low-income residents will likely not increase due to the exemption and that water prices in the city will likely be more stable (as the water reuse saves some freshwater resources of the city, which may lead to more stable water prices, benefitting particularly low-income residents). The more these implications were viewed as positive for low-income residents, the more was the policy perceived as fair for this group.Of the four policy implications for future generations living in San Francisco, two explained perceived fairness of the policy for this group and were both on average perceived as positive. These implications were that the policy leads to innovation in the wastewater treatment sector which will benefit future generations and that it leads to financial savings for the municipal utility (and thus the city), which may benefit future generations. The more positive these two implications were perceived, the fairer was the policy perceived for future generations living in San Francisco.The two implications of the policy for the local environment both explained perceived fairness for the environment. One was on average perceived as negative, namely the environmental risk in case of system failures as untreated or partially treated water used for irrigation may contaminate the environment. The more this implication was perceived as negative, the more was the policy perceived as unfair for the environment. The second implication was on average perceived as positive, namely that more water stays in the local and regional ecosystems since less water is taken from ground or surface water reservoirs. The more positive this implication was seen, the fairer was the policy perceived for the environment.Across all groups and entities, perceived fairness of the policy was explained both by implications perceived on average as positive and those perceived as negative. Of the 30 implications included in the study, 18 were on average rated as positive and 12 as negative, of which 11 and 7, respectively, significantly explained perceived fairness. The Fisher’s exact test (two-tailed, p = 1) indicated that neither the positive nor the negative implications were more likely to explain perceived fairness of the policy.Further, both implications related to financial aspects of the policy (n = 14) and those related to other aspects (n = 16) explained perceived fairness. Of these, 12 and 6, respectively, significantly explained perceived fairness. A Fisher’s exact test (two-tailed, p = 0.011) indicated that implications of the policy related to financial aspects were more likely to explain perceived fairness of the policy than non-financial ones.Perceived fairness for different groups and entities explaining policy acceptanceOverall, the policy mandating on-site systems was perceived as moderately acceptable (M = 4.81, SD = 1.26). A considerable amount of variance in acceptance was explained by perceived fairness, but only by fairness for three of the five societal groups included in the study. Most strongly, policy acceptance was explained by the perceived fairness for the city of San Francisco and its population, followed by the perceived fairness for future generations living in San Francisco and by the perceived fairness for owners of buildings with mandated on-site systems. With higher perceived fairness for these groups, the policy was evaluated as more acceptable, while it was perceived as less acceptable if it was perceived as less fair for these groups. Interestingly, the policy was on average perceived as rather fair for all groups and entities except for residents of buildings with mandated on-site systems, for whom the policy was on average perceived as slightly unfair (but close to the ‘neutral’ midpoint of the scale). However, perceived fairness for residents did not explain policy acceptance. For the bivariate correlations of all fairness variables and acceptance, see Table S7 in the Supplementary Information.DiscussionWith increasing water scarcity around the world, (on-site) wastewater reuse has become increasingly important2,4 and with it policies that regulate its implementation. Therefore, for a policy mandating on-site reuse in San Francisco, the present study investigated, (1) whether and to what extent different implications of the policy for five groups of society and one entity (the environment) explain perceived fairness of the policy for the specific group or entity, and (2) whether and to what extent the perceived fairness of the policy for each group or entity explains policy acceptance. The following groups and one entity were included in the study: residents of buildings with mandated on-site systems, owners of buildings with mandated on-site systems, the city of San Francisco and its population, low-income residents of San Francisco, future generations living in San Francisco, and the local and regional environment.Across all groups and the one entity, policy implications that were, on average, considered positive were equally likely to explain the perceived fairness of the policy as those that were considered negative. Thus, perceived costs, risks, and benefits arising from a policy all shape the public’s fairness evaluations. Interestingly, of all implications, those that were related to financial aspects (i.e. financial costs, risks, and savings, the creation of green jobs and cheaper technologies, stable rents and water prices) were more likely to explain perceived fairness than non-financial implications. Specifically, for all groups for which the policy has financial implications (i.e. all included groups but the environment), at least one of them explained perceived fairness. Thus, financial implications were central for fairness evaluations across all societal groups and not only for low-income groups. This is in line with studies on other environmental policies that found financial aspects to be relevant for perceived fairness or acceptance30,44,45,46. This finding implies that it might be particularly important in policymaking processes to consider the potential financial implications (both costs and benefits) of the policy for different groups of society. Financial burdens for some groups could be counterbalanced, for example by providing subsidies, lowering taxes, or reducing fees for public utility services. Moreover, freshwater could be priced appropriately (i.e. be less subsidised) to increase the financial benefit for those people reusing their wastewater29,47. Some of these measures are already in place in San Francisco, such as the option of reduced water and wastewater capacity charges for buildings that are covered by the mandate48 as well as a charge for the excess use of water beyond a designated threshold to promote the reuse of water49. Yet, it is essential that measures are in place ensuring that the financial burden is not shouldered entirely by the residents while the financial benefits remain with the building owners.Nevertheless, also six non-financial implications significantly explained perceived fairness. These can be grouped in three thematic pairs. One pair is related to a positive, sustainable image, both for residents and for owners of buildings with on-site systems. The more positive their image was perceived, the fairer was the policy evaluated for them. This matches findings of other studies that found increased acceptance of on-site systems41 and of other types of technologies50,51,52,53 if the technology use was associated with a better image or status gain. Additionally, this is in line with the ‘innovation diffusion theory’54 and the ‘costly signaling theory’55, both postulating that adopting (sustainable) innovations is associated with higher social status.The second thematic pair is related to the policy’s environmental impacts. Specifically, participants’ fairness evaluations for the environment were explained by the perceived environmental risk in case of system failures and the benefit that more water stays in the local ecosystems. This aligns with Kollmann et al.41, who reported acceptance of on-site systems to be higher, the more people perceived them as beneficial to the environment. Interestingly, the perceived fairness for future generations was not explained by perceived environmental benefits (i.e. saving water resources for future generations). Instead, only perceived financial benefits related to innovations in the local water and wastewater infrastructure explained fairness. Interestingly, also the third pair of non-financial implications that explained perceived fairness is related to San Francisco’s infrastructure. Specifically, the perceived risk of a reduced maintenance and water quality for residents of buildings with on-site reuse compared with users of the centralised system and the perceived benefit of an increased wastewater treatment capacity.These insights into non-financial implications explaining perceived fairness of a policy can be used for the design of communication strategies employed during the planning and implementation of policies on wastewater reuse. For example, the sustainable and innovative nature of the technologies, as well as the potential for positive publicity for builders and users could be emphasised. Moreover, risk communication strategies could address the perceived risk of a reduced maintenance and water quality for users of on-site systems by emphasising the safety of the treatment and reuse as well as by informing about monitoring and safety measures in place.Policy acceptance was strongly explained by perceived fairness for three different groups of society. This strong association is in line with previous studies on policies mandating water reuse25,29,56,57 and environmental policies in general9,23 and underlines the central role of fairness evaluations for policy design. Specifically, acceptance was explained by perceived fairness for three of the five societal groups included in the study, namely the city of San Francisco and its population, future generations living in San Francisco, and owners of buildings with mandated on-site systems. Acceptance was not explained by the perceived fairness for residents of buildings with mandated on-site systems, low-income residents of San Francisco, or the local and regional environment.The finding that the perceived fairness for future generations explains policy acceptance suggests that collective considerations impact people’s fairness evaluations of environmental policies. Similar results were found in a study on a policy mandating on-site reuse in India29 and a study on a transport pricing policy30. However, in both of these studies, fairness for future generations was assessed in combination with fairness for the environment. Thus, it is not possible to determine for which group or entity fairness explained policy acceptance. As in the present study policy acceptance was not explained by perceived fairness for the environment, it could be concluded that perceived fairness for future generations might be more relevant for policy acceptance than perceived fairness for the environment. Yet, more research on this question is needed. Consequently, future studies investigating the relation between perceived fairness and acceptance of environmental policies should consider separate fairness assessment for future generations and the environment.Fairness for current residents of San Francisco also explained policy acceptance, which further emphasises that people consider collective outcomes. Yet, this finding might also reflect participants’ consideration of their personal outcomes. As none of the participants in the survey were covered by the policy, the population of San Francisco is likely the group of society they most identify with, and which is most closely related to their personal outcome. This might also explain why this group’s fairness is the one most strongly related to policy acceptance, as it may reflect both collective and self-centred considerations, which have both been shown to be central for fairness and acceptance of environmental policies29,30,33,43,58.In addition, the findings indicate that people consider the impact a policy has on those most strongly affected by it. Specifically, the fairness for owners of the buildings with mandated on-site systems explained policy acceptance. Interestingly, however, the fairness for residents of these buildings did not explain policy acceptance, even though this group is also strongly affected. This contradicts the finding of Kollmann et al.29 that perceived fairness and acceptance of a policy mandating on-site reuse in India was explained by the outcome for residents. Potentially, having to use an on-site system is perceived as a bigger burden on residents in India, where the systems often do not function well, and the residents are at higher risk of an impaired water quality, while users of such systems in San Francisco are much less likely to face such issues. Indeed, while both positive and negative policy implications were perceived for residents, the policy was perceived neither as unfair nor as fair for residents (i.e. the average fairness perception was close to the scale midpoint).This could explain why, in San Francisco, fairness for residents did not explain policy acceptance.Lastly, we did not find perceived fairness for low-income residents of San Francisco to explain policy acceptance. This is particularly interesting as environmental justice (i.e. the equitable distribution of costs, risks, and benefits of environmental decisions among members of society, regardless of their race, ethnicity, gender, or socio-economic status) is a key argument in the debate on water resource allocation in California and is considered pivotal for the successful implementation of water innovations21,59 and for sustainable urban planning in general60,61,62. It is especially relevant given that California legislated the human right to water in 2012, acknowledging the fundamental right of all residents to have access to safe, clean, and affordable water63. Notably, the design of the policy mandating water reuse in San Francisco bears both benefits and costs for the city’s low-income residents as low-income housing projects are exempted. For that reason it might be possible that participants did not form a strong opinion regarding the policy’s impact on low-income residents, which might explain why it did not explain policy acceptance. In fact, the policy was, on average, perceived as moderately fair for low-income residents, and three out of four implications for this group were perceived as moderately positive, which might explain why fairness for low-income residents was not associated with policy acceptance. The finding also sheds light on the question raised in the introduction whether there might be a general pattern that marginalised groups of society might not be considered in the context of policies mandating on-site reuse (or technologies in general) or policies that cover only part of the population. While the present finding deviates from those of other studies on environmental policies30,32,33,34,35, it is in line with the one other study investigating a policy mandating on-site reuse for part of the population29. In that study, the outcome for low-income residents did also not explain perceived fairness or acceptance of the policy. Together, the two findings support the conclusion that for policies mandating on-site reuse (or technologies in general) or that cover only part of the population, low-income groups may be considered less than in the context of other environmental policies. This could either be because, compared with pricing policies, the implications are not ‘just’ financial but also behavioural. Alternatively, for policies that cover only part of the population, people may focus on the outcome for those covered, while additional implications for vulnerable groups may be overlooked. To investigate this further, future studies could investigate whether this pattern consists for policies covering only part of the population but in a different context, unrelated to on-site systems or other technologies.Taken together, the findings suggest that individuals evaluate the acceptability of a policy based on its perceived fairness for the collective (i.e. population of San Francisco and future generations) as well as for those directly and strongly affected (i.e. building owners). However, considerations of fairness do not necessarily extend to all affected or particularly vulnerable groups of society. Potentially, perceived fairness and policy acceptance are shaped less by a group’s marginalised or affected status per se and more by the specific impact of the policy on that group. Given that participants received information about specific implications of the policy for the different groups, it is likely that they integrated this information into their fairness judgements. Consequently, participants may have developed stronger fairness perceptions for some groups over others, which could explain why perceived fairness for certain marginalised or affected groups, such as low-income residents and residents of buildings with on-site reuse, did not explain policy acceptance.To the best of our knowledge, the present study is the first that investigated whether and how strongly the perceived fairness for different groups of society and the environment explains acceptance of a policy mandating on-site reuse and that investigates which policy implications explain whether the policy is perceived as (un)fair to the individual groups or the environment. It is also one of only few studies that investigated perceived fairness of a policy designed to support climate change adaptation instead of mitigation. Nevertheless, a few limitations need to be mentioned. First, while the range of policy implications included in the study covered an extensive range, it was likely not exhaustive. Moreover, it should also be borne in mind that the occurrence of some of the implications is uncertain, as they will, if at all, only manifest themselves over the next few years or decades. This has been communicated to participants, but we do not know whether and how they integrated the uncertainty in their evaluations. Second, the range of societal groups or entities selected for inclusion in this study, though more comprehensive than in previous studies, may be incomplete. These groups and entities were chosen based on previous literature29,30,64 and based on the conducted stakeholder interviews. Yet, it is possible that groups that were not considered in this study could explain part of the remaining variance in perceived fairness. Lastly, some limitations concerning the sample should be noted. The collected sample size was only sufficient to detect effects of at least small to medium size. Thus, it is possible that small effects were not detected. Furthermore, our study comprised only residents of San Francisco not covered by the policy. We know from Kollmann et al.29 that fairness and acceptance ratings can differ between people covered by the policy and those not covered. Thus, we cannot draw confident conclusions about the perceptions of residents of San Francisco covered by the policy. Moreover, the majority of study participants were highly educated and had a high income, while marginalised groups, such as residents with a lower income or education, were underrepresented. This might have been exacerbated by the exclusion of participants who failed to answer the knowledge questions and attention checks correctly. This restricts the validity of the conclusions that can be drawn from the study regarding the policy perceptions of these groups. Therefore, future studies should investigate the research questions among residents covered by the policy and, in particular, among those covered and with a low-income.Taken together, for the policy mandating on-site reuse in San Francisco, perceived fairness for the five groups of society and the environment was explained by several policy implications for the groups and the environment. This included both implications that were on average considered positive and implications that were on average considered negative by participants as well as implications related to financial aspects of the policy and implications related to non-financial aspects. This implies that all these different types of implications should be considered by policymakers when evaluating the fairness implications of policies related to on-site reuse. Moreover, policy acceptance was explained by the perceived fairness of the policy for the city of San Francisco and its population, future generations living in San Francisco, and owners of buildings with mandated on-site systems. This suggests that, overall, fairness for the collective as well as for those most affected is important to people when assessing policy acceptance. However, not necessarily the fairness for all people affected or for particularly vulnerable groups of society is considered.MethodsStudy location and policy designSan Francisco faces increasing water scarcity due to population growth, more stringent regulations about in-stream flows, and changing climatic conditions, such as increasing temperatures and reduced snowpack in the Sierra Nevada. In response to these issues, San Francisco mandated the installation and use of an on-site system for new construction projects in 201565. The latest version of the policy applies to construction projects of ≥100,000 gross ft² and mandates residential buildings to install an on-site system for the collection and treatment of the building’s greywater (i.e. wastewater from sinks, showers, washing machines, dishwashers etc. but not from toilets) as well as their condensate (e.g. from air conditioners). The recycled water has to be reused for the following non-potable purposes within the building or its premises: clothes washing, toilet flushing, and irrigation. Moreover, the full demand of water for these purposes must be met by the recycled water.Notably, low-income housing projects are exempted from the mandate. As a consequence, low-income construction projects are less expensive compared with conventional projects that fall under the mandate, as no on-site system has to be installed. The exemption aims at encouraging developers to build low-income housing. For a very limited number of low-income housing projects that voluntarily install on-site systems, San Francisco’s utility offers a funding scheme to offset costs.Procedure and sampleData were assessed through an online questionnaire, which was programmed with Unipark and distributed to the general public of San Francisco via the survey panel company Bilendi between April and June 2023. Participation took around 25 min and was financially compensated. All participants gave informed written consent prior to participation. The study protocol was approved by the institutional review boards of Eawag and the University of California, Berkeley [2021-08-14578] and was pre-registered on the Open Science Framework (OSF) on 04/27/23 before data collection started (osf.io/5c3z9). Prior to the data collection, semi-structured, virtual interviews were conducted between March and September 2022 with 12 San Francisco-based key stakeholders, including representatives from the water utility, the water quality control board, public advocacy groups, and the plumbers union as well as with developers of on-site systems, architects, and property managers. These interviews informed the design and content of the survey.Residents of San Francisco above the age of 18 were eligible for participation. Of the 1020 people who started the survey, 523 participants were screened out before the main part of the questionnaire because of incorrect answers to multiple-choice questions on the content of an information text on on-site systems and the policy in San Francisco (see section Questionnaire and measures). Of the remaining participants, 116 did not complete the questionnaire. During data cleaning, 205 participants were removed from the sample because of either of the following reasons: repeated participation (n = 28), speeding (defined as being faster than one third of the sample median; n = 56), failed attention checks (n = 55), straightlining (n = 27), or random answers to open questions (n = 39). Of the final sample (N = 176), 51.1% identified as female, 46.0% as male, and 0.6% (one person) as non-binary. On other person preferred to self-describe their gender, and 1.7% did not indicate their gender. The participants’ age ranged from 18 to 80 years (M = 48.29; SD = 17.59). All but three participants (98.3%) had completed high school, with 78.5% having completed tertiary education (Bachelor’s degree or higher). About half of the participants (54.8%) indicated having a yearly income of over $100,000. None of the participants reported living in a building with an on-site system. Compared with those who dropped out or were excluded during data cleaning, participants included in the analyses were significantly older (t(212.61) = 6.43, p < 0.001), were more likely to be female (χ2(4) = 10.21, p = 0.037), and had a lower education level (χ2(4) = 13.58, p = 0.009) and lower income (χ2(4) = 24.78, p < 0.001).A sensitivity power analysis66 was conducted to test for the smallest effect size detectable in the most demanding analysis conducted, namely a linear multiple regression with eight predictors given a power of 0.80 and an α = 0.0567. The omnibus F-test indicated a smallest detectable effect size of f2 = 0.09, corresponding to a small to medium effect68.Questionnaire and measuresAfter giving their consent to participate and indicating their sociodemographic data, participants read an informational text about on-site systems and the policy in San Francisco, followed by two multiple-choice questions on the content of the text (see ‘Supplementary Note 1’ for text and questions). As the following part of the questionnaire required a basic understanding of on-site reuse and the policy, only participants who answered the questions correctly could proceed with the questionnaire. They were given three trials to give the correct answers with the option to read the informational text in between. Those participants who could proceed were then presented with items assessing the valence of different policy implications for the five different groups of society and the environment as well as the perceived fairness of the policy for these groups and the environment. The items were assessed separately for each group or entity and in random order. Finally, participants’ acceptance of the policy was assessed. See ‘Supplementary Note 2’ for all items.The policy implications and the groups and entities of society included in the study were selected on the basis of existing literature29,30,64 and guidelines8,65 as well as the qualitative interviews with local key stakeholders, to ensure that the implications and groups were both comprehensive and relevant to the local context. For each group or entity of society, between two and nine implications were included in the questionnaire and presented in random order. For example, one implications for residents was ‘Residents have to bear the recurring costs of operation, monitoring, and maintenance of the systems. Moreover, it is likely that the initial costs of installation are passed on to the residents by the builders’. Participants were asked to rate each implication with regard to how positive or negative they are for the respective societal group or entity on a 7-point rating scale ranging from 1 (very negative) via 4 (neither negative nor positive) to 7 (very positive).To reduce complexity of the data and to avoid multicollinearity in the analyses, item mean scores were created for items that had a similar content, correlated highly and had acceptable internal consistency assessed with Spearman-Brown coefficients69. Specifically, means were calculated for the following items: implications for residents of buildings with mandated on-site systems with regard to (1) reliable and (2) unrestricted water supply in case of natural disasters or droughts (ρ = 0.67), implications for low-income residents of San Francisco with regard to the lack of (1) monetary and (2) (non)-monetary benefits of on-site reuse (ρ = 0.81) and, for the same group, with regard to the benefit of (1) stable rents and (2) water prices (ρ = 0.72). For the regression analyses, the means of these items were used.Directly after rating the policy implications for one of the societal groups or the environment, participants were instructed to rate the fairness of the policy for the respective group or entity when considering all of the stated implications for the specific group or entity. For example, the item for residents was ‘For residents of buildings with on-site reuse, the policy is overall…’, rated on a scale ranging from 1 (very unfair) via 4 (neither unfair nor fair) to 7 (very fair). The design of the item was adapted from previous studies on fairness and acceptance of environmental policies30,70.Policy acceptance was assessed with five semantic differential scales that were subsequently combined into one mean scale (α = 0.94). The participants were asked to rate the items ‘Overall, the policy is…’ on five scales ranging from 1 (…very unacceptable / negative / unnecessary / intolerable / useless) to 7 (…very acceptable / positive / necessary / tolerable / useful). The scale also included a midpoint, for example: 4 (…neither unacceptable nor acceptable). The scale was adapted from previous studies on acceptance of environmental technologies and policies29,71.Statistical analysesTo analyse which policy implications explain perceived fairness of the policy for different groups or entities, six multiple linear regression analyses were conducted for each of the five groups and one entity considered. Additionally, and across all implications and fairness ratings, two exploratory Fisher’s Exact Tests were conducted to analyse if perceived fairness is more likely to be explained by (1) either financial or non-financial implications and (2) either implications on average considered positive or those considered negative. Another multiple linear regression analysis was conducted to examine whether and to what extent perceived fairness of the policy for the five different societal groups and the environment explains policy acceptance. Owing to the non-normal distribution of the residuals, the analyses were conducted using bootstrap estimation with 10,000 replications. Significance was determined based on the bootstrapped 95% confidence intervals72. All analyses were conducted using IBM SPSS Statistics (Version 29)73.

    Data availability

    The datasets generated in this study are shared at: https://osf.io/pkv62/overview?view_only=0743ff496c79475ab322b28d62e6ac3b.
    Code availability

    The code used in this study is shared at: https://osf.io/pkv62/overview?view_only=0743ff496c79475ab322b28d62e6ac3b.
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    Resolving inherent constraints in eutrophication monitoring of small lakes using multi-source satellites and machine learning

    AbstractRemote sensing monitoring of small-lake eutrophication faces challenges such as sparse data, insufficient synergy of multi-source data, and limited model generalization performance. Hence, this study developed a scenario-aware modeling framework for the trophic level index (TLI) by integrating multi-source imagery data from Sentinel-2, GF-1, HJ-2, and PlanetScope, using Dongqian Lake in Zhejiang Province, China as the case study. The cross-sensor prediction accuracy was evaluated using algorithms such as CatBoost Regression (CBR), XGBoost Regression (XGBR), TabPFN Regression (TPFNR), and Linear Regression (LR). Meanwhile, the influence of input features was quantified by SHapley Additive exPlanations (SHAP). The main results found that : (1) Overall annual mean values of total nitrogen/total phosphorus ratio (TN/TP) and TLI were 22.13 and 37.36 ± 4.99, respectively, indicating a mesotrophic and phosphorus-limited state in Dongqian Lake. (2) TLI exhibited the strongest correlation with water color and algal spectral indices, including Normalized Difference Water Index (NDWI), Normalized Green–Red Difference Index (NGRDI), and Blue–Green Ratio (BGR). (3) CBR demonstrated the strongest cross-sensor generalization capability across different imagery, with only minor variations in prediction accuracy (ΔR ≈ 0.07–0.15). Feature attribution analysis identified NDWI, NGRDI, and BGR as primary contributing features for the CBR model. (4) Integrating high-frequency multi-source remote sensing imagery with 27 field surveys achieved seamless monitoring of the TLI. The spatial distribution of TLI showed distinct seasonal variations, with higher values observed in nearshore areas and lower values in the lake center. TLI values were relatively low in spring, but surged sharply and remained elevated in summer. This study provided a reference basis for detailed remote sensing monitoring and management of eutrophication in small lakes.

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    IntroductionLakes, as critical freshwater resources, play essential roles in sustaining the global hydrological cycle and ecological balance. They contribute significantly to biodiversity conservation, water supply regulation, and local climate moderation1,2,3. However, existing research has predominantly focused on large lakes, while the ecological value and localized regulatory functions of small lakes—defined as those with <100 km² surface area—have been largely overlooked4,5. Limited water exchange capacity and a higher susceptibility to external nutrient inputs make small lakes more prone to eutrophication, accompanied by lower self-purification potential and heightened ecological vulnerability6. Targets 6.3 and 6.6 of the United Nations Sustainable Development Goals emphasize the need for improving freshwater quality and protecting aquatic ecosystems7. In parallel, China’s “10-Point Water Plan” advocates for refined and localized waterbody management8. Accordingly, developing high-frequency, high-accuracy remote sensing monitoring systems tailored for small lakes is crucial for effective evaluation and adaptive regulation in water environment governance.The Trophic State Index (TSI) is a widely used indicator to assess lake trophic status9. However, it fails to fully reflect the nitrogen-phosphorus imbalance characteristics in Chinese lakes and reservoirs10. To address this, the TSI has been modified by incorporating total nitrogen and the permanganate index, thereby creating a more comprehensive evaluation index, the Trophic Level Index (TLI), which is better suited to the unique characteristics of water bodies in China11. The TLI integrates transparency (Secchi depth, SD), chemical oxygen demand by manganese (CODMn), total phosphorus (TP), total nitrogen (TN), and chlorophyll-a (Chla), offering a more accurate representation of eutrophication levels in lakes12.Recent remote sensing monitoring of lake eutrophication mainly used four methods: empirical models, semi-analytical models, machine learning methods, and deep learning methods13,14,15. Empirical models are structurally simple and cost-effective, widely used to rapidly estimate water quality parameters16. However, they are highly sensitive to changes in water optical properties and struggle to adapt to complex or highly heterogeneous lake environments17. Semi-analytical models enhance the physical interpretability of models by parameterizing water spectral reflectance mechanisms. However, they rely heavily on prior knowledge and parameter calibration, making them difficult to scale for regional applications18. Although deep learning methods have strong nonlinear fitting capabilities and excel in large-sample, high-dimensional remote sensing scenarios, they face challenges in small-lake remote sensing monitoring due to limited sample sizes, high overfitting risks, and a lack of interpretability19.In comparison, machine learning methods effectively explore nonlinear feature relationships under medium and small sample conditions. They offer strong modeling flexibility and good interpretability, providing an important tool for remote sensing water quality inversion20. In January 2025, Hollmann et al.21 published in Nature the TPFN (Tabular Prior-Data Fitted Network) model, combining Bayesian priors and neural network architecture, enabling efficient regression modeling under small sample conditions without hyperparameter adjustment, demonstrating strong cross-task generalization ability. However, its applicability in remote sensing with multi-source, multi-dimensional, and temporal data remains unverified. Conversely, eutrophication studies of small lakes based on remote sensing are still significantly constrained by insufficient data. Furthermore, most studies are limited by reliance on a single remote sensing platform, often beset by insufficient spatial resolution, infrequent coverage cycles, and inadequate temporal continuity. These studies rarely provide systematic comparative analyses spanning multiple remote sensing images and models, thus neglecting the relationship between model interpretability, imagery, and TLI22,23,24,25. These limitations hinder the high-frequency and dynamic monitoring of eutrophication processes in small lakes. Especially for lakes with limited areas and significant environmental changes, obtaining stable and continuous remote sensing data becomes a key obstacle to modeling.Consequently, this study proposed an scenario-aware modeling framework to address the challenges of monitoring eutrophication in small lakes. This framework combined multiple remote sensing imagery sources, including the Sentinel-2 multispectral instrument (S2), GF-1, HJ-2, and PlanetScope (PS). This method could improve the spatial and temporal continuity of monitoring, ensuring more comprehensive coverage of the study area. Multiple machine learning models were employed, including CatBoost regression (CBR), XGBoost regression (XGBR), TPFN regression (TPFNR), and linear regression (LR). A systematic comparison of these four models evaluated their accuracy and cross-sensor generalization capabilities under multi-source remote sensing imagery. Thus, the model’s features were quantitatively analysed using SHapley Additive exPlanations (SHAP) to develop a robust and universally applicable TLI inversion model.The specific objectives of this study are: (1) to enhance eutrophication monitoring capabilities for small lakes by integrating multi-sensor data in high-frequency field observations; (2) to investigate how the accuracy and generalization capabilities of different machine learning models vary across different remote sensing imagery and temporal dimensions; and (3) to determine the seasonal and spatial distribution characteristics of Dongqian Lake using TLI. This study provides the technical basis for developing high-resolution eutrophication monitoring systems for small lakes and offers a scientific foundation for formulating water quality management and ecological conservation policies.ResultsModel framework overviewTo assess the effectiveness and performance differences among various regression methods for predicting the trophic status of small lakes using different sensors, a scenario-aware modeling framework was developed. Representative machine learning models were selected: CBR, XGBR, TPFNR, and LR. Constructing the input dataset for remote sensing features involved combining features derived from various images with field sampling data. To assess the predictive performance and generalization capability of the models, model evaluation metrics were applied. Ultimately, the model demonstrating the highest accuracy and the lowest error across a range of remote sensing images was chosen for the final TLI inversion. The scenario-aware modeling inversion framework is illustrated in Fig. 1.Fig. 1: The scenario-aware modeling framework for estimating TLI.Step 1—collect optical imagery and field data (Chla, TP, TN, SD, CODMn); Step 2—preprocess/engineer features and select the optimal model; Step 3—apply atmospheric correction and lake-pixel extraction to estimate and map monthly TLI.Full size imageStatistical profile of in-situ water qualityThe overall mean values of water quality parameters for the entire year were: DO at 8.38 ± 1.38 mg·L−1, SD at 49.02 ± 14.85 cm, Chla at 38.79 ± 29.83 mg·m−3, CODMn at 4.27 ± 1.21 mg·L−1, the TN/TP at 22.13, and TLI at 37.36 ± 4.99, indicating that the lake was in a mesotrophic state (Table 2) with phosphorus limitation. Correlation analysis of the overall data for the year revealed that TLI had a strong significant correlation with Chla (r ≈ 0.85, p < 0.001) (Fig. 2); significant correlations with TP (r ≈ 0.75, p < 0.01) and TN (r ≈ 0.66, p < 0.01); and a significant correlation with CODMn (r ≈ 0.55, p < 0.05). In addition, TLI showed a significant negative correlation with SD (r ≈ −0.70, p < 0.01). However, TLI had no significant correlations with Tem (r ≈ 0.50), pH (r ≈ 0.35), and DO (r ≈ −0.30), all with p > 0.05. To analyze the distribution characteristics of water quality parameters across different seasons, we employed histograms, normal distribution fitting curves, and violin plots for visual representation. Furthermore, we applied the Shapiro-Wilk test to evaluate normality and to emphasize the statistical distribution differences of the parameters on a seasonal scale.Fig. 2: Correlations between TSI and water quality parameters.For clarity, correlation values below 0.1 are grayed out. The correlations are significant at levels of 0.05 (*), 0.01 (**), and 0.001 (***) (two-tailed).Full size imageTLI seasonal variations and related parameters demonstrated significant differences, with the distribution of each parameter significantly deviating from normal distribution (p < 0.0001)(Fig. 3). In spring, DQ Lake was generally in a mesotrophic state (TN/TP > 22.6, indicating phosphorus-limitation), with a slow increase in Chla, a relatively low TLI, moderate CODMn, higher DO, and good water transparency (SD), indicating stable water quality. In summer, the lake reached its peak eutrophication level, with TN/TP ranging from 9 to 22.6 (dual nitrogen and phosphorus limitation), and some areas dropping below 9 (nitrogen limitation). Chla increased rapidly, and TLI reached a light eutrophic state, with CODMn peaking at its highest level of the year. DO showed significant fluctuations throughout the day, and water transparency noticeably decreased. In autumn, the lake began its recovery process after eutrophication, with TN/TP levels remaining between 9 and 22.6, indicating dual nitrogen and phosphorus limitation. Although Chla decreased, it remained at a high level, while TLI exhibited a gradual decline. CODMn levels stayed elevated, low oxygen conditions continued, and water transparency remained poor, which indicates a slow recovery process. During winter, the lake reached a state of stabilization, with the TN/TP ratio exceeding 22.6, suggesting phosphorus limitation. Chla hit its lowest levels of the year, and TLI decreased in parallel. CODMn reached its lowest level of the year, while DO saw a significant increase. Additionally, water transparency improved markedly, leading to enhanced water quality. These seasonal differences validated the overall trends observed in the annual correlation analysis, further clarifying the seasonal patterns of water quality parameter variations.Fig. 3: The seasonal statistical distribution of water quality parameters.Panels: Seasonal violin plots for TLI, Chla, TN/TP ratio, CODMn, and DO. Black curve: Fitted normal distribution. Annotation (below each panel): Shapiro-Wilk test results (statistic and p-value) for data pooled across all seasons, indicating significant deviation from normality (p < 0.01).Full size imageSelecting remote sensing features for TLI modelingTo identify the key features suitable for TLI remote sensing modeling, this study analyzed the relationship between features from the S2, GF-1, HJ-2, and PS datasets and TLI, using significance (p < 0.05) and r as criteria (Fig. 4). After outlier removal based on IQR method, 279, 78, 124, and 68 valid sampling points were retained for S2, GF, HJ, and PS data, respectively. In the S2 dataset, NDWI (r = 0.47), NGRDI (r = 0.46), and BGR (r = 0.46) showed strong significant correlations with TLI (p < 0.01). NDCI and CI, showing significant correlations (r = 0.31, p < 0.05), were included in the feature selection process. In the GF-1 dataset, NDWI (r = 0.64, p < 0.01) made the greatest contribution, followed by BGR and NGRDI (r = 0.43, p < 0.01). Although NDTI and some bands (B1-B4) exhibited negative correlations, they were also selected due to their statistical significance. In the HJ-2 dataset, there was a significant correlation between BGR and NGRDI (r = 0.44, p < 0.01), and bands B1, B2, and B5 satisfied the selection criteria (r ≈ 0.31–0.32, p < 0.05). In the PS dataset, the features B8, NDTI, NDCI, CI, GCI, and EVI also met the selection criteria (r ≈ 0.28–0.31, p < 0.05), while other features fell short of the required threshold.Fig. 4: The r values between TLI and remote sensing variables derived from S2, GF-1, HJ-2, and PS datasets.Panels (S2, GF-1, HJ-2, and PS) represent results from four satellite datasets. Color bars: Indicate the magnitude and direction of the correlation coefficient (r). Asterisks: Denote statistical significance: *p < 0.05, **p < 0.01, ***p < 0.001 (two-tailed tests).Full size imagePerformance assessment of models with multi-source dataEstimating TLI across different satellite sensorsTable 1 displays the predictive performance of the regression models CBR, XGBR, LR, and TPFNR across various remote sensing datasets. Overall, the CBR model outperformed other models in terms of both prediction accuracy and stability. In the GF-1 and HJ-2 datasets, the CBR model achieved R of 0.94 and 0.83, respectively, with NRMSE of 0.13 and 0.16 and biases close to zero, indicating high prediction accuracy and robustness. In the S2 dataset, the correlation coefficients of the CBR and TPFNR models were similar (R = 0.80 and 0.81, respectively). However, CBR had significantly lower errors (NRMSE = 0.14, Bias = 0.15) than TPFNR (NRMSE = 0.16, Bias = 0.26), showing better prediction stability and accuracy. The high spatial resolution of PS remote sensing imagery led to increased computational errors in the model, which caused a decrease in the prediction accuracy of the CBR model (R = 0.64, NRMSE = 0.19, Bias = −0.09). However, it still surpassed other models, demonstrating its superior adaptability to the data. The XGBR model demonstrated high fitting accuracy on the training set (with R values generally close to 1). However, its generalization performance on the test set was insufficient, especially in the HJ-2 dataset, where the bias was large (Bias = 0.66). In the PS dataset, the NRMSE (0.21) was also higher than the CBR model, suggesting the presence of overfitting. The TPFNR model performed well in some datasets (e.g., S2 and GF-1) but demanded a significantly higher computational cost than other models (with training times reaching 68.5 seconds). Additionally, its performance in the PS dataset dropped substantially (R = 0.50, NRMSE = 0.24), indicating limitations in model complexity and generalization ability. The LR model recorded consistently poor performance across all datasets, with the highest error observed in the PS dataset (NRMSE = 0.24) and the lowest prediction accuracy (R = 0.47), demonstrating its inadequate ability to capture complex nonlinear relationships in remote sensing data.Table 1 Performance comparison of models for TLI retrieval based on S2, GF-1, HJ-2, and PS datasetsFull size tableTo further assess the generalization capability of the models, a comparison was conducted between the predicted values of the CBR, XGBR, TPFNR, and LR models and observed values (Fig. 5). The analysis reveals significant differences among the models across various sensor types and sampling time points. Of the models evaluated, the CBR model exhibited the best generalization performance, with its predicted values closely aligning with the 1:1 reference line. For both S2 and GF-1 datasets, CBR achieved high consistency with observations, with no notable overestimation or underestimation tendencies. In the HJ-2 dataset, slight underestimation was observed at higher TLI, while minor overestimation was predicted at lower TLI. In the PS dataset, prediction scatter increased, particularly underestimating at high TLI values. In contrast, XGBR and TPFNR models displayed value-dependent error characteristics, particularly under HJ-2 and PS datasets. Significant underestimation was noted at high values, while predictions at low values deviated substantially from the 1:1 line, indicating potential overfitting. The LR model consistently underperformed across all datasets, with pronounced underestimation at high values and highly scattered results at low values, particularly for GF-1 and PS, leading to considerably higher error levels than other modelsFig. 5: The scatter plots depict the performance of various models (CBR, XGBR, TPFNR, and LR) in retrieving TLI from the S2, GF-1, HJ-2, and PS datasets, covering both training and test sets.The graphs featuring red boxes highlight the top-performing models for each sensor.Full size imageAnalyzing the contribution of SHAP-based featureTo clarify the contributions of individual remote sensing features to model predictions, SHAP was utilized to evaluate feature importance across four models: CBR, XGBR, TPFNR, and LR. (Fig. 6). The results showed both consistent patterns and notable differences in feature importance, attributed to variations in the spectral characteristics of the data and the structures of the models.Fig. 6: Feature importance based on SHAP values for models (CBR, XGBR, TPFNR, and LR) in predicting TLI across S2, GF-1, HJ-2, and PS datasets.Graphs with red boxes indicate the best-performing model for each sensor.Full size imageThe CBR model consistently relied on vegetation indices and composite spectral features across all sensor types. In the S2 dataset, the primary predictors of CBR were identified as NDWI, BGR, NGRDI, and NDVI. For the GF-1 dataset, CBR mainly depended on BGR, B3, and NGRDI. In the HJ-2 dataset, the important features included NGRDI, BGR, NDTI, and B1. In the PS dataset, NDVI and NDTI emerged as the most influential features for CBR. The XGBR model integrated diverse spectral features to effectively utilize multi-dimensional spectral information. Within the S2 dataset, shortwave bands (B7, B6, B2) and index-based features (NDVI, NDCI, NGRDI) played a significant role in the XGBR model. The GF-1-based XGBR model highlighted a combination of BGR, NDWI, and NDTI, while the HJ-2 XGBR model focused on B6, NDTI, and B1. For the PS dataset, NDVI and NDTI were key variables in the XGBR model. In contrast, the TPFNR model showed strong feature selectivity, heavily relying on mid- to near-infrared bands and specific indices. In the S2 dataset, the key features for the TPFNR model included B7, B6, NDCI, and B11. For the GF-1 dataset, NGRDI and BGR were emphasized in the TPFNR model. The preferred features for the TPFNR model in the HJ-2 dataset included NGRDI, BGR, NDTI, and B1. In the PS dataset, the TPFNR model prioritized GCI, NDCI, and CI.The LR model showed limited feature utilization across all datasets, with predictive power largely concentrated on a small subset of bands. In the HJ-2 dataset, B3, B5, and B2 accounted for most of the LR model’s variance, underscoring the linear model’s restricted capacity to capture complex spectral patterns.Assessing TLI spatial and temporal variabilityThe integration of multi-source imagery effectively compensates for the lack of effective observations from individual sensors on specific sampling dates, ensuring temporal continuity between sampling and image matching, and enabling near-seamless spatiotemporal monitoring. Preprocessed images from S2, GF-1, HJ-2, and PS were fed into the CBR model to generate spatial distribution maps of TLI values for DQ Lake during various periods in 2023 (Fig. 7). Overall, TLI exhibited distinct seasonal dynamics. In early January, TLI values across the lake ranged from 32 to 33.5 and remained relatively low in early spring. From late spring to mid-May, TLI values experienced a significant increase, initially rising from 32 to 39 and later reaching levels between 44 and 49. During the summer months of June and July, TLI values continued to exhibit an upward trend, attaining peak values ranging from 34 to 36.25 and 38 to 46, with the exception of a sharp decline on June 14, 2023. However, by late August, TLI rebounded to higher levels, fluctuating between 37 and 46.Fig. 7: Spatial distribution of TLI retrieved from S2, GF-1, HJ-2 and PS imagery in 2023.The figure comprises a series of TLI maps organized by satellite source and acquisition date, covering scenes from S2, GF-1, HJ-2, and PS sensors throughout 2023. The color bar indicates the TLI value range.Full size imageAt the spatial scale, TLI displayed a consistent pattern of spatial heterogeneity, characterized by a distinct gradient: the values were higher in nearshore areas and decreased toward the lake center. The northern bay and southwestern coastal areas have consistently been zones with high TLI values at different points in time. In contrast, the central and southern open-water areas exhibited persistently low TLI levels. Notably, TLI values reached a maximum of 44 to 46 in nearshore zones during the summer, with occasional high-value regions emerging in shallow or shoreline areas on specific dates.DiscussionIn practical remote sensing applications, inversion models generally require data processing from multiple satellite platforms, which exhibit significant discrepancies in sensor specifications, spatial resolution, spectral configurations, and imaging conditions. Therefore, ensuring consistent generalization performance across diverse sensor configurations is paramount for reliable water quality prediction. The results indicate that all models experience a decline in prediction performance when applied across different sensors. However, the extent of degradation and stability present significant variations among the models. Among the evaluated models, CBR demonstrates higher stability and generalization capability across sensors by partitioning the feature space into numerous regions and performing weighted aggregation of similar samples within the same region. The CBR model demonstrates the most consistent performance, exhibiting a reduction in the R ranging from 0.07 to 0.15. It is noteworthy that the model shows relatively high prediction accuracy when transitioning from higher-resolution data, such as GF-1, to lower-resolution data, such as HJ-2, and from S2 to PS. SHAP-based feature importance analysis confirms the model’s consistent reliance on vegetation-related spectral indices (e.g., NDWI, NGRDI, CI, and BGR) across all sensor datasets, indicating strong adaptability to heterogeneous data sources. Conversely, the XGBR model displays notable variations in cross-sensor prediction performance, particularly in scenarios characterized by substantial disparities in spatial resolution. This sensitivity is likely attributable to the model’s high dependence on input feature consistency. The TPFNR model demonstrates notable performance instability across sensors and incurred substantially higher computational costs due to its structural complexity. This finding indicates that the quality and consistency of the data significantly influence the model’s generalization. The LR model demonstrates the least effective cross-sensor generalization capacity. The linear structure of the model proves inadequate in capturing the nonlinear spectral variability inherent in remote sensing data, thereby limiting its predictive capability across different sensor platforms. Consequently, models such as CBR, which demonstrate superior generalization performance, should be prioritized in practical applications of remote sensing-based water quality monitoring. For models showing weaker generalization capacity (e.g., XGBR and TPFNR), it is imperative to exercise additional caution to mitigate the uncertainties introduced by sensor-specific differences.The variations in model prediction performance can be attributed to various factors, including model structure, preprocessing strategies, image quality, and observed values. Meanwhile, the variations in spectral features of different sensors are primarily influenced by their physical characteristics:

    (1)

    Differences in band availability and spectral response functions are key factors. For example, S2 and PS are associated with red edge bands centered around 705‒740 nm, whereas GF-1 and HJ-2 satellites lack dedicated red edge coverage. Consequently, red edge dependent indices, NDCI and CI, used in this study are standard configurations for S2 and PS data, but require alternatives, NDCI* for NDVI-type and CI* for NIR/R-1 on HJ-2.

    (2)

    Due to differences in spatial resolution and shoreline proximity effects, subpixel mixing is more obvious in narrow waterways and along bright shorelines. As a result, indices constructed from red/green bands (e.g., NGRDI, BGR) exhibit increased sensitivity to these non-water signals, leading to reduced robustness and greater susceptibility to positive bias near shorelines.

    (3)

    Different sensors utilize distinct atmospheric correction workflows. S2 adopt the 6S model, GF-1 and HJ-2 use the FLAASH model, and PS uses the DSF model, respectively. Residual errors persist in the visible–near-infrared (VIS/NIR) bands, due to different assumptions and inputs among the methods, such as incomplete removal of haze path radiation or effects from shoreline proximity.

    Figure 7 shows the spatial distribution of the TLI consistently register higher values in the nearshore zone. This phenomenon may be attributed partly to human and natural factors, such as tributary nutrient inputs, enclosed bays, and extensive aquatic grass beds. However, some higher values may be induced by optical interference factors, including shallow-water reflectance, shoreline shading effects, and signals from aquatic vegetation that enhance surface reflectance. Although high-frequency sampling can enhance reliability, some factors may reduce the accuracy, such as nearshore substrate heterogeneity, anthropogenic disturbance, and hydrodynamic influence. Consequently, during the image pre-processing stage, we first excluded nearshore areas by applying a water body mask. After inversion, high-value nearshore regions were designated as “hotspots” and analyzed for their rationality based on spatial patterns and temporal consistency.Despite the comprehensive evaluation of model performance and cross-sensor generalization using multi-source remote sensing data, some limitations have remained. They include the relatively short temporal span of available data, the limited spatial coverage of in-situ sampling, and the lack of fully integrated multi-source strategies. To further advance eutrophication monitoring in small lakes, future studies should focus on extending the temporal span of remote sensing and in-situ datasets, refining data preprocessing and sampling protocols, and promoting the integrated use of domestic and international satellite resources, thereby improving model accuracy, robustness, and generalizability.MethodsStudy areaDongqian (hereinafter referred to as “DQ”) Lake is a typical small coastal plain lake with shallow depths. Human activities have significantly impacted it, making it an ideal case study for investigating eutrophication in small lakes. Situated in the eastern part of Yinzhou District in Ningbo City, Zhejiang Province, China, it is the largest natural freshwater lake in the province (Fig. 8)26. The elongated lake is 8.5 km long (north-south axis) and 4.5 km wide, with 22 km² water surface area and 45 km shoreline. With an average depth of 2.2 m, the water volume amounts to 33.9 million m3. DQ Lake is classified as a lagoon type. The region has a subtropical monsoon climate, with mild and humid conditions throughout the year, an average annual temperature of 15.4 °C27. The lake is pivotal in providing water for drinking and agricultural irrigation for Ningbo, which is important for regional water resource security and ecosystem services.Fig. 8: Geographic location of the study area.a Location of Ningbo City in Zhejiang Province, China; b Land use and land cover around DQ lake56; c Topography around DQ lake and distribution of in-situ water sampling sites.Full size imageIn-situ water quality assessmentsThe water-quality dataset was provided by the Ningbo Ecological and Environmental Bureau, covering the sampling period from January to October 2023. A total of 26 fixed monitoring points were established around the lake (Fig. 8c). Monitoring was conducted 27 times, resulting in 702 samples. Sampling was performed four times per month from May to September, and once per month for the remaining months. Temperature (Tem), Dissolved Oxygen (DO), pH and Secchi Depth (SD) were measured in situ. Chemical Oxygen Demand-Manganese (CODMn), Total Phosphorus (TP), Total Nitrogen (TN) and Chlorophyll-a (Chla) were analyzed in the laboratory. Table 2 shows the descriptive statistics for the water quality parameters.Table 2 Statistical information on the water quality parameters in DQ LakeFull size tableSatellite data acquisition and processingThis study’s satellite imagery primarily comprised S2 data retrieved via Google Earth Engine (GEE), supplemented by HJ-2 and GF-1 imagery obtained from the China Centre for Resources Satellite Data and Application (CRESDA), and PS imagery sourced from Planet Labs (Supplementary Table S1 for details)28. The following processing steps were implemented:

    (1)

    All imagery was matched within a ± 2-day window of the sampling date for each field survey. When multiple images were available, the one with the closest temporal synchronization, cloud-free condition, and highest spatial resolution was selected (Supplementary Table S2 for details).

    (2)

    Reliable auxiliary data on aerosols and water vapor were available, and where the distance between land and water bodies was moderate, the 6S reflectance model was utilized with Sentinel-2 satellite data29. MODTRAN effectively handled proximity effects and path radiation. In cases where bright coastlines or densely populated urban areas caused notable proximity effects and haze, the FLAASH model was applied for GF-1 and HJ-2 satellites30. Dark spectral fitting (DSF) was proven to be highly effective in both inland and turbid waters without relying on external aerosol data. Therefore, the DSF albedo model was applied to PlanetScope satellite imagery31,32.

    (3)

    Water bodies were extracted using the Normalized Difference Water Index (NDWI)33.

    (4)

    To minimize interference from shoreline mixing, image data showing a water proportion less than 0.8 within a 3 × 3 pixel window were excluded.

    Methodology for TLI estimationOutlier detection was performed using the Interquartile Range (IQR) method, which can ensure the inversion accuracy of the model. Each variable’s first (Q1) and third (Q3) quartiles were calculated, and IQR was defined as IQR = Q3 − Q134. Observations falling outside the range of [Q1 − 1.5 × IQR, Q3 + 1.5 × IQR] were identified as outliers and removed35.TLI was constructed with Chla (mg m−3) as the core parameter, along with SD (cm), TP (mg L−1), TN (mg L−1), and CODMn (mg L−1). TLI was calculated based on the national standard36:$$TLI(Chla)=10(2.5+1.086,{text{ln}},Chla)$$
    (1)
    $$TLI(TP)=10(9.436+1.624,{text{ln}},TP)$$
    (2)
    $$TLI(TN)=10(5.453+1.694,{text{ln}},TN)$$
    (3)
    $$TLI(SD)=10(5.118-1.94,{text{ln}},SD)$$
    (4)
    $$TLI(CO{D}_{Mn})=10(0.109+2.66,{text{ln}},CO{D}_{Mn})$$
    (5)
    $$Wj=frac{{r}_{ij}^{2}}{mathop{sum }_{j=1}^{m}{r}_{ij}^{2}}$$
    (6)
    $$TLI(varSigma )=mathop{sum }limits_{Wj}cdot TLI(j)$$
    (7)
    $$mathrm{TLI}(varSigma )=0.2663mathrm{TLI}(mathrm{Chla}+0.1879mathrm{TLI}(mathrm{TP})+0.1790mathrm{TLI}(mathrm{TN})+0.1834mathrm{TLI}(mathrm{SD})+0.1834mathrm{TLI}({mathrm{COD}}_{mathrm{Mn}}))$$
    (8)
    n the formula, TLI(j) represents the composite index for parameter j, with the associated weight Wj. The rij refers to the correlation coefficient between the reference Chla and each parameter j (r2Chla = 1, r2TP = 0.7056, r2TN = 0.6724, r2SD = 0.6889, r2CODMn = 0.6889)37. The final TLI value is calculated as the weighted average of the individual indices. The trophic state classification standard for Chinese lakes (reservoirs) is shown in Table 338:Table 3 Trophic classification criteria for lakes and reservoirs in ChinaFull size tableIn lake ecosystems, the TN/TP ratio is commonly utilized to identify the primary limiting nutrient factor, indicating the relative availability of nitrogen or phosphorus during phytoplankton growth39. A TN/TP ratio below 9 suggests that the water may be nitrogen-limited, which can promote the growth of nitrogen-fixing cyanobacteria. On the other hand, when the TN/TP ratio exceeds 22.6, phosphorus limitation may occur, hindering phytoplankton growth. If the TN/TP ratio falls between these two thresholds, the water body may experience simultaneous nitrogen and phosphorus limitations40. Analyzing the TN/TP ratio is crucial for understanding how changes in lake nutrient profiles relate to risks of eutrophication.To enhance the accuracy of remote sensing data inversion models and to improve their sensitivity to spectral variations, this study developed a multidimensional spectral feature dataset. This dataset was derived from atmospheric-corrected surface reflectance products obtained from various sources of remote sensing imagery. It is important to note that differences exist in the band configurations across different remote sensing platforms (see Supplement Table S3 for details). The dataset mainly includes: (1) Water Optical Characteristics indicators, such as the Normalized Difference Turbidity Index (NDTI) and Normalized Suspended Sediment Index (NDSSI), which primarily reflect changes in water turbidity and suspended particulate matter41,42. (2) Algal Biological Response indicators, including the Phytoplankton Algae Index (FAI), Chlorophyll Index (CI), Normalized Chlorophyll-a Index (NDCI), and Green Chlorophyll Index (GCI), which characterize the dynamic changes in algal pigment content and chlorophyll-a in water bodies43,44,45. (3) Key Bands and Ratio Combinations, such as the Blue/Green Ratio (BGR) and various single-band reflectance indices (B1−B12), which help capture water absorption characteristics and spectral curve shape variations46. During the feature selection process, the Pearson correlation coefficient (r) was utilized to assess the relationship between each variable and TLI. Variables that demonstrated insufficient significance were eliminated based on a two-tailed t-test (p < 0.05)47.Machine learning models for TLI retrieval. CBR is an improved gradient-boosting method that utilizes ordered target encoding and oblivious decision trees48. It efficiently handles heterogeneous features without complex preprocessing, offering strong generalization capabilities and interpretability. XGBR offers advantages such as regularization control, missing value handling, and parallel computation, making it effective for modeling nonlinear relationships49. Particularly well-suited for high-dimensional and noisy data, it is widely applied in remote sensing inversion and feature recognition tasks. TPFNR is based on a Prior-Function Network structure, leveraging meta-learning from millions of tabular tasks for model initialization. It does not require manual parameter tuning and can achieve rapid convergence and robust predictions in small-sample scenarios. The classic linear model LR has a simple modeling mechanism and low computational cost, making it suitable for modeling linear or approximately linear relationships50.The dataset was randomly split into training (70%) and testing (30%) sets (Rhodes et al. 2023). For CBR and XGBR, hyperparameter tuning was performed using grid search combined with k-fold cross-validation. The key adjusted parameters were learning rate, L2 regularization (CBR), and gamma (XGBR)51. TPFNR and LR were automated modeling methods requiring no manual parameter configuration. After model training, the SHapley Additive exPlanations (SHAP) method was used to explain feature importance in each model, enhancing model interpretability52. The TPFNR model was implemented using the Python TabPFN package. CBR and XGBR were executed via CatBoost (v1.2.3) and XGBoost (v2.0.3), respectively, while the LR model was supported by scikit-learn (v1.3.1). The model parameter configurations utilized for TLI prediction in multi-sensor experiments are detailed in Table 4.Table 4 Model parameter configurations used for TLI prediction in multi-sensor experimentsFull size tableTo systematically evaluate the predictive performance of the models in TLI inversion, this study utilizes four key metrics: Pearson correlation coefficient (r), Normalized Root Mean Square Error (NRMSE), Bias, and Unbiased Mean Absolute Percentage Difference (UMAP)53,54. These metrics comprehensively reflect model performance from the perspectives of correlation, margin of error, and systematic bias, ensuring the thoroughness and robustness of the evaluation55. The following equations were used to calculate:$$NRMSE=frac{sqrt{frac{1}{n}{sum }_{i=1}^{n}{(xp-xt)}^{2}}}{frac{1}{n}{sum }_{i=1}^{n}xt}cdot 100 %$$
    (9)
    $$Bias={x}_{p}-{x}_{t}$$
    (10)
    $$UMAP=frac{1}{n}mathop{sum }limits_{i=1}^{n}frac{|U{D}_{i}|}{{x}_{t}}$$
    (11)
    Where xp represents the predicted values obtained from the model and xt denotes the observed values. The Unbiased Difference (UD) is the residuals resulting from the LR fitting between the predicted and observed values. UD reflects how accurately the model’s predictions align with the observations.

    Data availability

    The datasets generated and analyzed during the current study are not publicly available due to the data access restrictions imposed by the funding research project, but are available from the corresponding author on reasonable request.
    Code availability

    The underlying code for this study is not publicly available but may be made available to qualified researchers on reasonable request from the corresponding author.
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    Reprints and permissionsAbout this articleCite this articleSi, W., Chen, Z., Jim, C.Y. et al. Resolving inherent constraints in eutrophication monitoring of small lakes using multi-source satellites and machine learning.
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    Hydrological drought trends and seasonality in selected Polish catchments between 1993 and 2022 using a threshold based approach

    AbstractOngoing climate change and Land Use Change (LUC) in Europe are altering the frequency and seasonality of low-flow events. Increases in mean annual temperature, shifts in catchment land use, and reduced retention capacity contribute to intensifying streamflow deficits. This study examines long-term trends and seasonality of hydrological droughts in selected catchments of the Central European Lowland in Poland over hydrological years 1993–2022. Low-flow events were identified and characterised using tools in the R programming environment. Trends were assessed with the Mann–Kendall (MK) test and Sen’s slope estimator, and persistence was evaluated with the Hurst exponent. Significant correlations (p < 0.05) were found between the Total Number of Drought Events (TNDE) and LUC, particularly for events lasting ≥ 7 days at Q90 threshold, with the strongest effects in the Skora, Luciąża, Widawka and Grabia rivers. A marked increase in summer and autumn low-flow events was recorded after 2002, alongside rising Seasonal Number of Drought Days (SNDD) and Seasonal Cumulative Deficit Volume (SCDV), reflecting growing water deficits. The persistence of these trends suggests continued streamflow decline. The findings can inform the identification of catchments vulnerable to worsening summer water conditions.

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    03 February 2024

    Introduction Drought is a complex and recurrent natural hazard that affects ecosystems and societies across diverse temporal and spatial scales1,2,3. It has both environmental and socio-economic impacts, making it one of the most impactful natural disasters. Droughts result from precipitation deficits of considerable magnitude, persisting over extended periods. Because these deficits propagate through various stages of the hydrological cycle, four main types of droughts are commonly distinguished: meteorological, agricultural, hydrological, and socio-economic droughts4.The occurrence and severity of drought are strongly influenced by two key factors: climate change and LUC5. Moreover, the combined effects of natural processes and human activity are the primary drivers of changes in climatic extremes and the resulting shifts in hydrological extremes6. According to Beran and Rodier7 and the Flow Regimes from International Experimental and Network Data (FRIEND) research programme4, drought is understood as a sustained, regionally extensive period of below-average natural water availability, whether in precipitation, runoff, or groundwater. Human activities and the resultant global warming have been identified as the main causes of the increasing frequency and intensity of drought events8. Globally, drought frequency and intensity are expected to increase, particularly in the Northern Hemisphere, with regions, such as the USA, China, India, South Korea, Vietnam, and Europe projected to experience worsening conditions5,9,10,11,12,13,14.Both climate change and LUC are known to pose risks to water resources15. Climate change has intensified extreme hydro-climatic events, and it is likely to continue increasing their frequency and intensity in the near future16,17. In recent decades, global warming has been well documented, with the mean global temperature rising by 1.19 ± 0.12 °C above pre-industrial levels18,19, and further increases are anticipated20. These shifts are expected to have a huge impact on hydrological processes21.At present, one of the principal tasks of hydrology is to identify regions vulnerable to the adverse effects of prolonged drought arising from climatic, hydrological, and hydrogeological conditions. To this end, it is essential to establish uniform criteria for distinguishing hydrological drought periods within long-term data series and to quantify this phenomenon22. Poland is among the regions highly vulnerable to severe hydrological droughts23. The areas most susceptible to atmospheric drought in Poland are the Central Polish Lowlands and the western part of the Pomeranian Lakeland24,25. Records of river low-flows have long been kept, with the earliest references to exceptionally low water levels in Polish territories dating back to 988, 1121, 1332, and 1473 26. Severe hydrological droughts affecting the whole of Poland occurred in the first half of the 1950 s, the early 1960 s, the mid-1980s, the early 1990 s, the mid-2000s, and the early 2010 s, indicating that Poland experiences significant water shortages on average every 10 to 15 years; however, prolonged low-flows with substantial streamflow deficits across all rivers simultaneously have been a rare phenomenon23. Poland has experienced several severe hydrological droughts over the last few decades, notably in 1992–1993, 2006 and 2008. In both 2012 and 2015, record low water levels of the River Vistula in the capital, Warsaw, were widely reported in the Polish media27. The year 2022 was considered particularly catastrophic in Europe in terms of hydrological drought. During the summer of 2022, Central and Southern Europe experienced an extreme drought, characterised by exceptionally low soil moisture and river water levels, severely affecting sectors across many countries28.Processes occurring on the Earth’s surface influence the intensity of extreme events, such as heat waves and droughts. Land transformation in mid-latitude regions has significantly increased the frequency of hot, dry summers29. Human-induced climate change leads to an accelerated water cycle and a more variable climate, which in turn results in greater fluctuations in river discharge30. Hydrological droughts, characterised by gradual onset and prolonged duration, result in substantial human and economic losses, with wide-ranging impacts31,32. In recent decades, severe droughts have affected nearly every continent33,34,35.Notwithstanding the findings from earlier studies it remains unclear which causes of extreme hydrological events are decisive. Although the impact of climate change on droughts is well-documented, the effects of LUC are less understood. Recent research has identified climate change as the primary driver behind the projected increase in drought frequency36,37. However, there is limited information on hydrological changes in smaller catchments, which can substantially influence entire river basins. In smaller catchments, low-flow periods are more evenly distributed over time, with hydrological droughts in the 1960 s and 1970 s affecting mountain rivers more severely, while those of the 1980 s and 1990 s had a greater impact on lowland catchments23. Small rivers play a crucial role in regulating basin-wide water flow and quality, acting as natural buffers for groundwater recharge. Their condition directly affects the stability of the wider basin, contributing to nutrient cycling, supporting biodiversity, and serving as critical water sources for agricultural and ecological systems. In lowland river basins dominated by agriculture, water availability for plant growth is particularly important28. In lowland areas, the duration of low-flows depends on basin water resources and the direction of water management operations, yet severe low-flow periods do not necessarily correspond proportionally to preceding ones23. Consequently, alterations in small river systems may exert substantial influence on the hydrological stability and functioning of the entire basin38,39.There is a wealth of information in the literature on hydrological drought prediction for future years based on hydrological models40,41. For monitoring droughts, numerous typical drought indices have been developed based on meteorological and hydrological variables to assess the magnitude, duration, severity, and spatial extent of droughts. One example is the Hurst exponent (H), first introduced by Hurst42 in his analysis of Nile River floods. Since then, this exponent has been applied in many scientific disciplines. Recent research on Hurst behavior in climate and hydrology is reported, for example, by O’Connell43 and Adarsh and Priya44. In this study, we used the Hurst exponent to evaluate the persistence of drought trends and their potential continuation in the study area.However, there is a lack of comprehensive studies, focused on small lowland catchments that exert a substantial influence on larger basins.This study aims to address that knowledge gap through a case study of several rivers located in the Polish lowlands (Fig. 1). Societal welfare and economic development in a region largely depend on effective integrated water resources management at catchment scale, as well as the implementation of drought monitoring and preparedness strategies. Accordingly, there is a growing need to further analyse hydrological droughts and their characteristics.Fig. 1Location of selected catchments.Full size imageIn this paper, we present a comprehensive drought analysis that involves identifying drought events and characterising their properties, including duration, intensity, cumulative deficit volume, and spatial extent. We also examine indicators related to drought seasonality and apply the peaks-over threshold (POT) method. Over the years, many indices have been proposed based on factors, such as the nature of the water deficit, temporal resolution, regionalisation, or standardisation. The most widely used approach for identifying drought periods is the POT method, in which a drought is defined as a period when river flow falls below a specified threshold (Qt). This method, known as the threshold approach, was introduced by Yevjevich in the United States45 and by Zielińska in Poland46. In this study, the POT method is applied to selected cross-sections, where low-flow events are analysed using three thresholds that correspond to different drought types. It is assumed that Qt = Qp, meaning the threshold flow corresponds to a selected percentile of flow exceedance. The most commonly used thresholds are Q70 and Q9047,48,49, while Tallaksen and van Lanen4 recommend using thresholds from Q70 to Q95. In Poland, a commonly recommended methodology includes three thresholds: Q70, Q90, and Q9550. Moderate drought is defined by the Q70 flow, which serves as a warning level; Q90 defines deep drought, considered a state of emergency; and Q95 represents extreme drought, typically classified as natural disasters. Considering these thresholds, the study aims to comprehensively analyse water deficits under all three variants and to evaluate the effects of land use changes within the analysed catchments. An additional objective is to calculate the Hurst exponent in order to evaluate the persistence of flow patterns and identify potential future trends in water availability.ResultsCatchmentsThe MK test was performed for 15 analysed rivers. The MK test indicated no statistically significant trend for the Nurzec River). Additionally, an increasing trend was identified for the Niesób River. Based on these findings, both rivers were excluded from further analysis. The Sen’s slope estimator confirmed the increasing trend for the Niesób River. In contrast, the remaining rivers exhibited negative slope values, indicating a decline in river flows over time. Notably, the Skora River showed the lowest Sen’s slope value, suggesting the most pronounced negative trend in water flow among the rivers studied (Table 1).Table 1 Hydrographical characteristics of selected rivers and gauging stations and significance and direction of trends in time series (1993–2022) based on the MK test. The direction of the trend is indicated by the sign of the sen’s slope estimator (+ or −).Full size tableBased on the 2018 CORINE Land Cover data (CLC), agricultural land dominates the analysed catchments, covering on average more than 64% of the area (Fig. 2). The Sama River catchment has the highest proportion of agricultural land at 80.6%. In contrast, the Miała River catchment is predominantly forested, with forested land accounting for 86.7% of its area. The catchment with the highest percentage of anthropogenic land use is the Widawka River catchment, where built-up areas represent 13.3% of the total surface (see Supplementary Fig. S1 online). An analysis of land cover changes between 2000 and 2018 shows that all catchments underwent measurable transformation. During this period, the percentage of anthropogenic land increased in all catchments, with an average gain of 2.58% points (pp). The largest increase was observed in the Skora River catchment, where the share of built-up areas rose by 7.3 pp at the expense of agricultural and forest land. Notable changes were also recorded in the Sama, Luciąża, Grabia, and Widawka catchments, where the share of anthropogenic land increased by 3.9–4.3 pp.Fig. 2Land use (LU) in 2000 (A) and 2018 (B), and LUC (C) for selected catchments. Legend: 1 – artificial surfaces; 2 – agricultural areas; 3 – forest and semi-natural areas; 4 – wetlands; 5 – water bodies.Full size imageThe analysed catchments were also classified according to soil types. The predominant soils are podzolic (PZ), which are highly permeable and cover an average of 36% of the area, and luvisols (LV), which are poorly permeable and account for approximately 38% of the area. The Miała River catchment lies almost entirely (97%) within areas dominated by PZ, indicating high permeability. In contrast, the Skora River catchment is primarily composed of LV (see Supplementary Fig. S2 online).Streamflow droughtThe analysis of hydrological data for selected rivers during the hydrological years 1999–2022 allowed for the assessment of total drought duration for periods of at least 7 days, classified into three levels: moderate (Q70), deep (Q90), and extreme (Q95) (Table 2).The rivers with the highest number of flow drought events (126, 115, and 112) are: the Ślęza River, the Skora River, and the Prosna River, respectively. The pattern of fewer, drought periods in rivers (i.e. Sama River and Bawół River) where TNDE was less than 50 for Q70 and lasted at least 7 days, can be attributed to the geological morphology (dominance of well-drained soils, such as FL, PZ and CM). In the case of the Ślęza River where the highest TNDE values were recorded for Q70 flow and minimum 7-day duration, extending the drought period to minimum 30 days resulted in only 28 events with flow limit value Q70, 8 events with limit value Q90, and 4 events with Q95. However, in rivers where low TNDE values were observed (for the threshold value of Q70 and duration of ≥ 7 days), such as the Grabia River and the Sama River, extending the minimum low-flow duration to ≥ 30 days led to higher TNDE values across all threshold flow variants compared to the other analysed rivers. Analysis of drought events lasting at least 20 and 30 days reveals less frequent but more persistent periods of streamflow deficit, which is of key importance for assessing the risk of prolonged water shortages in water management. (Tables 2 and 3). This suggests that in catchments with a high TNDE, drought episodes tend to be short and frequent, while in catchments with fewer events, the episodes are generally longer and more persistent, which indicates that drought periods are rare but long-lasting. Furthermore, as drought duration extends, the Mean Cumulative Deficit Volume (MCDV) increases across all rivers, with the highest MCDV observed in the Sama River during a moderate drought (Q70) lasting at least 7 days. The highest Mean Cumulative Deficit Volume (MCDV) was observed for the Ślęza River under the Q70 threshold for drought durations of at least 30 days (Table 4).Table 2 Hydrological values of the TNDE for the Q70, Q90, Q95 threshold levels at gauging stations on selected rivers (1993–2022).Full size tableTable 3 Hydrological values of the MDD for the Q70, Q90, Q95 threshold levels at gauging stations on selected rivers (1993–2022).Full size tableTable 4 Hydrological values of the MCDV for the Q70, Q90, and Q95 threshold levels at gauging stations on selected rivers (1993–2022).Full size tableA Student’s t-test revealed a statistically significant correlation between MDD and MCDV (r = 0.6521, p < 0.001), indicating a strong positive relationship. This analysis, performed collectively across all three examined percentiles, confirms an association between longer drought durations and greater cumulative water deficits. These findings indicate a heightened risk of water shortages and the occurrence of long-term droughts in the region (Fig. 3). Additionally, correlation analyses conducted separately for each percentile confirmed the overall observed relationship. The results demonstrated consistency across all cases, with statistically significant correlations between MDD and MCDV observed for each percentile. The correlation coefficients were r = 0.62 for Q70 (p < 0.05), r = 0.51 for Q90 (p < 0.05), and r = 0.48 for Q95 (p < 0.05).Fig. 3Correlation between the analysed parameters: TNDE, MDD, and MCDV (total for thresholds Q70, Q90, Q95).Full size imageFig. 4Heatmap with hierarchical clustering showing links between hydrological values and land use change (LUC) for selected rivers, based on drought durations of: A ≥ 7, B ≥ 10, C ≥ 20, and D ≥ 30 days.Full size imageThe analysis of heat maps depicting the frequency of low-flow events of varying durations, in conjunction with land cover change data, enabled the identification of spatiotemporal patterns from a hydrological perspective (Fig. 4). The cluster analysis of standardised results allowed the identification of several groups of rivers based hydrological and LUC values, depending on drought duration. For droughts lasting at least 7 days, a group was identified including the Sama, Mogilnica, and Bawół rivers, which were characterised by the highest values of MDD_Q90, MDD_Q70, and MCDV_Q70. A second group comprised the Luciąża, Skora, Ślęza, and Barycz rivers, which exhibited the highest values of TNDE_Q70, TNDE_Q90, and TNDE_Q95, with the Skora River additionally showing a high value of the LUC_1 parameter. This suggests that land use changes, such as urbanisation and land surface sealing, considerably reduce infiltration and retention capacity, thereby increasing the frequency of short-term low-flow events. Extending the drought duration to at least 10 days, and then to 20 days, revealed that the Sama, Mogilnica, and Bawół rivers continued to exhibit high MDD_Q70 values, and for droughts of at least 20 days, high TNDE_Q95 values were also observed. For droughts lasting at least 30 days, the Skora River stood out with the highest MCDV_Q90, MCDV_Q95, and LUC_1 values, while the Sama, Mogilnica, and Bawół rivers continued to exhibit high TNDE_Q90, TNDE_Q95, and MDD_Q70 values. These findings imply that the most pronounced changes in the Skora River’s flow regime are associated with short-duration droughts, which are more susceptible to catchment-scale modifications affecting infiltration and water retention. Furthermore, the clustering of rivers into correlation classes shifted with increasing low-flow duration, suggesting that catchment responses vary not only in intensity but also in temporal behavior, reflecting spatial heterogeneity in their reaction to changing precipitation and flow conditions.The conducted Principal Component Analysis (PCA) explained 60.7% of the total variance associated with LUC and the analysed hydroclimatic indicators (TNDE, MDD, MCDV) for low-flow events lasting at least 7 days. This threshold was selected as it marks the shortest period indicative of hydrological drought onset. The first principal component accounted for 44.8% of the variability, while the second explained 15.9%. A strong negative correlation was observed between the increase in anthropogenic land cover (LUC_1) and the decrease in agricultural land area (LUC_2), indicating that urban expansion has occurred largely at the expense of arable land. Furthermore, LUC_1 showed a positive correlation with the TNDE_90 indicator, suggesting that the growth of built-up areas may contribute to an increased number of days affected by severe low-flow conditions, based on the 90th percentile threshold. This relationship was particularly evident in the catchments of the Skora, Luciąża, Widawka, and Grabia rivers, which have experienced considerable spatial transformation in recent years, marked by an increased proportion of built-up areas (Fig. 5).Fig. 5Principal Component Analysis (PCA) of hydrological parameter values and land use change for catchments.Full size imageThe PCA performed separately for the years 2000 and 2018 revealed substantial changes in the LU_1 indicator (representing anthropogenic land cover) over the 18-year period. The results clearly show that by 2018, LU_1 became positively correlated with TNDE, indicating that the increase in human-modified land areas had begun to strongly influence the number of low-flow days. In contrast, in the 2000 analysis, LU_1 was a neutral factor with respect to TNDE (see Supplementary Fig. S3 online) This shift confirms the findings from the previously generated heat map, which also highlighted the growing impact of anthropogenic pressure on hydrological drought conditions.A detailed seasonal drought analysis was conducted by dividing the year into four seasons (spring, summer, autumn, and winter) to capture temporal variability in low-flow events. Three hydrological drought indicators were calculated: Seasonal Number of Drought Days (SNDD), Seasonal Number of Drought Events (SNDE), and Seasonal Cumulative Deficit Volume (SCDV).The analysis indicates a clear predominance of drought in summer and autumn across all percentile variants (Fig. 6). For the Q70 threshold, the mean SNDD reached 53 days in summer and 34 days in autumn. Under the Q95 threshold, which indicates extreme drought, these values dropped to 12 days in summer and just 1 day in autumn, suggesting that summer is the most drought-prone season in the studied region. Correlation analysis between the drought indices revealed statistically significant, strong relationships: SNDD and SNDE (r = 0.802, p < 0.05), and SNDD and SCDV (r = 0.702, p < 0.05). The strong correlation between SNDD and SNDE indicates that longer drought periods are associated with a greater number of drought events. Meanwhile, the positive correlation between SNDD and SCDV reflects an increasing water deficit in streamflow volumes during prolonged low-flow conditions, intensifying hydrological stress across catchments.Fig. 6Average values of SDCV, SNDD and SNDE indices in the hydrological years 1993–2022 for different thresholds: (A) Q70, (B) Q90, (C) Q95, and in different seasons.Full size imageThe Hurst exponent values for the analysed rivers ranged from 0.6844 to 0.8663, indicating strong autocorrelation and the persistence of trends in the flow time series (Table 5). This suggests that changes observed in successive years tend to follow the same directional pattern, pointing to the system’s so-called “hydrological memory”. The highest Hurst exponent values were recorded for the Główna and Miała rivers, suggesting that in these rivers, water flows are particularly influenced by past hydrological conditions and exhibit a high level of long-term predictability.Table 5 Hurst exponent for the analysed rivers, based on data from the multi-year period 1993–2022.Full size tableDiscussionPeriods of low river flow have a considerable impact on ecosystems, the economy, and human life. Climate change is contributing to increasing challenges related to water availability, leading to a growing water deficit on a global scale. Trnka et al.51 highlight a rise in the frequency, duration, and severity of droughts in Central Europe, directly attributable to climate change. In addition, land use changes, such as urbanisation, deforestation, and agricultural intensification, disrupt the natural water cycle, reduce infiltration, and increase surface runoff, further exacerbating water scarcity52. The duration of low-flow periods is strongly influenced by water management activities in catchments subjected to intense anthropogenic pressure. One example of this is the dewatering of open-pit mines, which has caused low-flow conditions to persist for extended periods in catchments affected by depression cones53. At the global level, projected climate changes is expected to intensify the hydrological cycle, increasing the frequency and risk of hydrological extremes such as droughts and floods54. Monitoring drought conditions is therefore essential for mitigation, with the threshold level method55 being among the most widely applied approaches.In recent years, Central and Southern Europe have experienced extreme drought characterised by exceptionally low soil moisture and river water levels. These events have had serious impacts across multiple sectors in several countries. Initially driven by a pronounced rainfall deficit, the meteorological drought eventually developed into low river flows with widespread effects28. By the middle of the century, both the frequency and intensity of heatwaves and droughts are projected to increase across most of Europe. This trend is already evident between 2000 and 2023, during which eight years recorded above-average drought-affected areas, with five of them within the last decade alone. The recurrence and growing severity of these events over the past 24 years suggest that a decline in drought-impacted areas by 2030 is unlikely56. A particularly pressing issue in Poland is the uneven distribution and variability of water resources across both time and space57. Recent changes in the atmospheric climate have led to a variety of impacts, among which alterations in the water balance due to a marked increase in field evaporation are the most pronounced. Since 1988, Greater Poland has been increasingly affected by water deficits, posing a direct threat to the supply needs of several key economic sectors essential to the region’s sustainability. The most severe precipitation deficits are observed in the central part of the country, which lies within Poland’s lowland region. As in many other areas of continental Europe, relatively low annual precipitation combined with rising temperatures has further exacerbated these pressures58. Drought is an inherent feature of the Polish climate, generally persisting for several weeks and affecting large areas. Its consequences can span multiple sectors. Even when the immediate impacts appear limited in scale or significance, they can be long-lasting and critical across various domains. Owing to its effects on the environment, economy, and society, increasing attention has been directed towards research and practical measures aimed at assessing, monitoring, and forecasting drought22.The threshold method applied to daily hydrographs from thirteen representative catchments in Poland enabled the identification of key characteristics of streamflow droughts over the past 30 years. This approach, which relies on daily flow variations rather than a fixed annual threshold, allows for a more accurate assessment of hydrological conditions, especially in regions with pronounced seasonality where different types of droughts may occur due to precipitation, temperature, or both. A similar method for defining drought periods was adopted by Rivera et al.59 and Seneviratnen et al.60. The analysis of the variability of regional average indicators in the hydrological years 1993–2022 indicates that drought conditions were particularly severe in 2003, based on SNDD and SCDV for the 70th and 90th percentiles. Studies by Laaha et al.61 and Sutanto and Van Lanen62 also indicate that 2003 was a pivotal year, recognised as one of the most important drought years in Europe. The average TNDE value for the Q70 threshold and a minimum drought duration of 7 days is 78 for the analysed catchments (Sama River), while MDD ranges from 20.99 to 68.80 days under the same input parameters. The longest low-flow period in 2003, at the Q70 threshold, was observed in the Bawół River, lasting continuously for 201 days (from 23 May to 9 December). In the Sama River, however, the longest low-flow period observed lasted 118 days (from 5 June to 30 September). The average MDD for Q70 and a minimum duration of 7 days is 42 days, whereas for a minimum duration of 30 days it increases to 77 days. The study by Sutanto and Van Lanen62 states that minor droughts are the cause of frequent occurrences of threshold-based droughts, both with variable thresholds (VTD) and fixed thresholds (FTD), in European rivers.In this study, MDD does not exceed 100 days at Q95 with the duration of the low flow being at least 30 days. By contrast, research by Tomaszewski and Kubiak-Wójcicka23 indicated that the average low-flow duration, at a 95% probability of non-exceedance, in Polish rivers is 162 days. The spatial variation of this parameter generally corresponds to the distribution of water resources. Their study analyzed low-flow conditions at 17 gauging stations across Poland between 1951 and 2015 and showed that, in most cases, the maximum duration did not exceed 250 days, although in some cases it exceeded 500 days23. Research on the number of low-flow days (NDLF) and long-term changes in river discharge was conducted by Wrzesiński et al.63, who analysed flows from 140 gauging stations located on 96 rivers across Poland for the period 1951–2020, as well as for two sub-periods: 1951–1988 and 1988–2020, representing the periods before and after climate change. Their study revealed changes in river discharge, with most gauging stations recording a decrease of approximately 5–15% (in some cases statistically significant). Furthermore, over the entire study period, statistically significant declining trends in NDLF were evident in the eastern and southern parts of the Vistula basin. After 1988, however, this pattern shifted, with NDLF showing an increasing trend, except for a few rivers in southern Poland, where statistically significant decreases were still observed.The dynamics of low-flow episode progression are characterized by high seasonal and multiannual variability23.In the present study, a seasonal analysis of drought in the rivers under investigation revealed a high incidence of the phenomenon during summer and autumn. At the Q70 threshold, the mean SNDD reached 53 days in summer and 34 days in autumn. Other authors also highlight that the autumn and summer months are the main periods determining severe hydrological drought. According to Tomaszewski and Kozek64, severe and extensive hydrological droughts usually occur during summer and early autumn (July–October), with June serving as a transitional month that can markedly prolong drought duration in years characterised by dry springs and summers. Moreover, in the majority of both upland and lowland rivers, the peak period of hydrological drought typically falls in August. Somorowska25, on the other hand, notes that droughts of varying intensity occur in Poland during both the winter and summer halves of the year. Her analysis indicates that the most extreme drought occurred in August 2015, lasting three months and affecting 47% of the country’s area.Given that these catchments span diverse geographical regions (mountains, lowlands, and coastal areas), it can be suggested that extremely severe hydrological droughts are influenced not only by hydrometeorological conditions but also by local catchment-specific factors, arising from a combination of natural processes and human activities65. In recent years, land cover in catchments has undergone noteworthy changes both in Poland and globally. These transformations influence various catchment processes, including water retention and river flow. Changes in land cover are driven by multiple environmental factors, such as water circulation, landscape quality, ecosystems66, and the hydrological regime67. Urbanisation alters catchments through the expansion of impermeable surfaces and modifications to drainage systems, which accelerate runoff and limit infiltration. This disruption of the hydrological cycle not only reduces groundwater recharge but also increases flood risk. The regional hydrological cycle, a key factor in flood management, has been significantly affected by both rising urbanisation and climate change, potentially leading to more extreme precipitation events and severe hydrological disasters in the near future68. An increase in impervious surface area reduces the retention capacity of catchments and contributes to local temperature rise69. Several studies also indicate that deforestation-related land cover changes are linked to an increased frequency of hydrological extremes70,71. For instance, rapid changes in the Prądnik catchment resulted in a 75% increase in water loss72. Intensifying catchment transformations contribute to the increasing severity of droughts, underscoring the importance of forecasting their future development and implementing effective mitigation strategies. There is an urgent need for sustainable water management solutions that preserve natural retention capacity, protect ecosystems, and support climate change adaptation. In this context, analysing long-term memory in hydrological processes using the Hurst exponent proves to be a valuable approach, as it enables assessment of trend persistence and potential future changes in runoff regime and water resources43,44. In the analysed area, the Hurst exponent for all studied rivers exceeded 0.5, with an average value of 0.78. This indicates the persistence of current hydrological trends and highlights the index’s value in strategic planning and water resource management. In their study, Millen and Beard72 applied the Hurst exponent to the Burdekin River in Australia, where the estimated value was 0.7527. As with the rivers analysed in this study, a value above 0.5 indicated that all hydrological variables are likely to follow persistent trends in the future. Tatli73 also used the Hurst exponent to examine drought conditions in Turkey, observing values close to 1 in areas vulnerable to future droughts. These findings support the assumption that long-range memory in large-scale climate systems and/or teleconnections plays a key role in explaining drought occurrence. Given that global warming is one of the major aspects of climate change, it further intensifies drought conditions. Similarly, studies conducted in Mongolia confirmed the reliability of the Hurst exponent in predicting drought trends, with a forecasting accuracy of up to 91.7%. In that study, the average Hurst value of the SPEI time series from 1980 to 2014 was 0.533, indicating that future drought trends are generally consistent with the present state74. The Hurst exponent can be a highly practical indicator for informing water and drought management policies. It has also been applied in flood risk prediction, demonstrating its versatility in hydrological assessments. To address the increasing challenge of hydrological drought, it is crucial to implement adaptive measures tailored to the specific characteristics of each region. This approach is essential for mitigating water deficits, preventing soil degradation, and ensuring reliable water availability for agriculture. Simultaneously, strengthening regional capacity and raising public awareness of climate change impacts on water resources are vital. Only comprehensive, locally adapted management of limited water resources can ensure their long-term and sustainable use73.The results of this study should be interpreted with caution, as they primarily reflect conditions in lowland rivers and may not be directly applicable to mountainous catchments. Future research could extend the analysis to include additional catchments and longer time series, which would allow for a more comprehensive understanding of regional drought patterns.MethodsThe study area encompasses 15 rivers situated in Poland’s lowland regions (Fig. 1), with catchment areas ranging from 176.01 km² to 726.28 km². These predominantly agricultural regions are highly dependent on sufficient water availability. Daily discharge data were obtained from 15 gauging stations, sourced from the hydrological database of the Institute of Meteorology and Water Management (IMGW). Figure 1 illustrates the spatial distribution of the selected gauging stations along with major Polish rivers. The 15 time series were selected based on data quality, spatial coverage, and the duration of available records. Table 1 lists the gauging stations and provides key details for each.Spatial dataThe boundaries of river basins and catchments, as well as vector layers of rivers, were obtained from the IIaPGW database (Second Update of River Basin Management Plans), available at https://apgw.gov.pl/. Vector data on land use for the years 2000 and 2018 were sourced from the Copernicus Land Monitoring Service (https://land.copernicus.eu). Vector data on soil types were obtained from the Harmonized World Soil Database (HWSD), version 2.0, available at https://data.isric.org/.Data analysisDaily flow data from 15 hydrological gauging stations, spanning a 30-year period (hydrological years 1993–2022), were analysed using statistical methods, specifically the MK test and Sen’s slope estimator. A 30-year period was selected as it represents the minimum standard recommended by the World Meteorological Organization (WMO) for calculating hydrological drought characteristics and for comparing hydrological data75. The MK test is a non-parametric, rank-based test used to detect monotonic trends within time series data. Given the often skewed distributions of hydrometeorological data, the MK test is particularly well suited for this purpose, as is the Sen’s slope estimator. Sen’s slope estimator was used to quantify the magnitude of the trend detected by the MK test, allowing for an assessment of the rate at which a given variable, such as river discharge, changes over the 30-year period. Failure to meet the significance of the MK test (p ≤ 0.05), together with the absence of a decreasing trend in the analysed flows, excludes the gauging station from further analyses. Both Sen’s slope and the MK test were performed in the R programming environment.Streamflow drought definitionTo identify streamflow drought events, we applied the threshold approach, originally proposed by Yevjevich45 and widely adopted in subsequent hydrological studies23,31,60,61,64. Threshold levels, also referred to as truncation levels, were derived from Severity-Duration-Frequency (SDF) curves at flow values equaled or exceeded 70%, 90%, and 95% (Q70, Q90, and Q95) of the time (≥ 7, ≥ 10, ≥ 20, and ≥ 30 days) respectively, calculated based on daily flow variations (For example SDF curves, see Supplementary Fig. S4 online). These thresholds represent normal, severe, and extreme water flow drought conditions, respectively. This approach enables the identification of both short-term and multi-year droughts4.A key advantage of this method, compared to standardised drought indices, is its ability to quantify deficit magnitude, which is an important factor for water resource management31. In this framework, a drought event begins when water flow falls below the threshold and ends when it rises above it.Hydrological indices, such as TNDE, MDD, MCDV, were calculated for different variants in the R programming environment, using the hydroTSM, remotes, dplyr packages. Subsequently, they were subjected to correlation analysis and visualised as a heat map combined with hierarchical clustering.To examine the interrelationships among the hydrological indicators (TNDE, MCDV, MDD) and land use changes, principal component analysis (PCA) was applied across all study catchments. Subsequently, PCA was also employed to investigate the association between hydrological indicators and land use. This analysis was conducted for two reference years, 2000 and 2018, across the entire set of catchments. Changes in LUC shares are expressed in pp. All graphs of PCA were created using the ggplot package in R studio.To show the seasonal variability of low-flow events, the SNDD, SNDE, and SCDV values ​​were calculated in the R programming environment using the dplyr and lubridate packages. Graphs in R were created using the ggplot package.Based on 30 years of daily data, the Hurst exponent was calculated to predict the future drought trend in the analysed catchments. The Hurst exponent was used to characterise the long-term memory of a time series, providing an indication of the persistence of prevailing trends or patterns. A Hurst exponent greater than 0.5 indicates strong persistence, whereas a value below 0.5 indicates anti-persistence76. The Hurst Index was calculated using the hursexp() function in the pracma R package.ConclusionsThe analysis of hydrological droughts in lowland catchments of Central Poland during the hydrological years 1993–2022 revealed significant changes in flow regimes (p < 0.001). Thirteen out of fifteen analysed rivers exhibited a decreasing trend in streamflow, indicating an intensifying water deficit in the studied region. The most pronounced drought occurrences were observed in the catchments of the Skora, Ślęza, and Prosna rivers, where the number of events exceeded 100 over the 30-year period. Land use changes, particularly the increase in anthropogenic areas at the expense of agricultural and forest lands, have a notable impact on the intensification of short-term drought episodes. This phenomenon was particularly evident in the Skora River catchment. The seasonality analysis confirmed that drought events are concentrated in the summer and autumn months, posing a major challenge for water resource management, especially in agriculture and the protection of aquatic ecosystems. Although long-duration droughts (≥ 20 and ≥ 30 days) occurred relatively infrequently, their characteristics, especially the high volume of MCDV, suggest increasing intensity and potentially longer periods of extreme drought events. High Hurst exponent values for the analysed rivers (0.68–0.87) indicate strong autocorrelation and persistence of observed trends, suggesting that unfavorable hydrological conditions are likely to persist in the future. The findings underscore the urgent need for integrated water resource management strategies at the local level, with particular attention to areas most vulnerable to drought, as small rivers play a crucial role in water retention and in maintaining the overall water balance of entire river basins. This methodological approach not only provides valuable insights into drought occurrence but also offers a transferable framework for similar hydrological assessments in other regions. In light of the observed changes, continued research is essential to better understand the impacts of climate and land use transformations on the hydrology of small catchments and their contribution to the functioning of entire river systems.

    Data availability

    All data are publicly available from the Institute of Meteorology and Water Management – National Research Institute (IMGW-PIB) in Poland, via the website https://danepubliczne.imgw.pl/or can be requested by contacting the corresponding author, Ewelina Janicka-Kubiak.
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    Download referencesAcknowledgementsThis work was supported by discipline of science environmental engineering, mining and energy in Poznań University of Life Sciences (Poland) as the research program „Innovator plus”, no. 04/2024/INN-PLUS.FundingThe research program “Innovator plus”, no. 04/2024/INN-PLUS. Poznań University of Life Sciences (Poland).Author informationAuthors and AffiliationsDepartment of Land Improvement, Environmental Development and Spatial Management, Poznań University of Life Sciences, Piątkowska 94, Poznań, 60-649, PolandEwelina Janicka-KubiakAuthorsEwelina Janicka-KubiakView author publicationsSearch author on:PubMed Google ScholarContributionsConceptualization, E.J.-K.; methodology, E.J.-K. software, E.J.-K.; formal analysis, E.J.-K.; investigation, E.J.-K. resources, E.J.-K.; data curation, E.J.-K.; writing—original draft preparation, E.J.-K.; writing—review and editing, E.J.-K.; visualization, E.J.-K.; supervision E.J.-K.; project administration, E.J.-K.Corresponding authorCorrespondence to
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    KeywordsStreamflow droughtHydrological valuesTemporal variabilityLand use changeHurst exponent More