s41467-025-67125-5.pdf

Article https://doi.org/10.1038/s41467-025-67125-5 Current protected areas provide limited beneﬁts for European river biodiversity James S. Sinclair 1 , Rachel Stubbington 2,E l l e nA .R .W e l t i3, Jukka Aroviita4,N a t h a nJ .B a k e r5, Miguel Cañedo-Argüelles6, Zoltán Csabai 7,8, David Cunillera-Montcusí 9,10,S a m iD o m i s c h11, Martial Ferréol12, Mathieu Floury12,13, Marie Anne Eurie Forio14, P e t e rL .M .G o e t h a l s14, Alexia M. González-Ferreras 15, Kaisa-Leena Huttunen16,17, Richard K. Johnson 18, Lenka Kuglerová19, Aitor Larrañaga 20,T i m oM u o t k a17, Riku Paavola 21,22,P e t rP ařil 23, Jes J. Rasmussen 24,R a l fB .S c h ä f e r25, Rudy Vannevel26,G á b o rV á r b í r ó27, Martin Wilkes28 & Peter Haase 1,29 Protected areas are a principal conservation tool for addressing biodiversity loss. Such protection is especially needed in freshwaters, given their greater biodiversity losses compared to terrestrial and marine ecosystems. However, broad-scale evaluations of protected area effectiveness for freshwater biodi- versity are lacking. Here, we provide a continental-scale analysis of the rela- tionship between protected areas and freshwater biodiversity using 1,754 river invertebrate community time series sampled between 1986 and 2022 across t e nE u r o p e a nc o u n t r i e s .P r o t e c t e da r e a sp r i m a r i l yb e n eﬁted poor-quality communities (indicative of higher human impacts) that were protected, or that gained protection, across a substantialp r o p o r t i o no ft h e i ru p s t r e a mc a t c h - ment. Protection had little to no inﬂuence on moderate- and high-quality communities, although high-qualitycommunities potentially provide less scope for effect. Our results reveal the overall limited effectiveness of current protected areas for freshwater biodiversity, likely because they are typically designed and managed to achieve terrestrial conservation goals. Broadly improving effectiveness for freshwater biodiversity requires catchment-scale management approaches involving larger and more continuous upstream protection, and efforts to address remaining stressors. These approaches would also beneﬁt connected terrestrial and coastal ecosystems, thus gen- erally helping bend the curve of global biodiversity loss. Biodiversity is in crisis owing to human-induced global change 1– 3. Extensive actions have been implemented to address these losses, including legislation and agreements to expand the cover of protected areas (PAs), such as the EU Habitats Directive (92/43/EEC) and the Kunming-Montreal Global Biodiversity Framework, which sets a target of 30% global PA coverage by 2030 4. PAs restrict or reduce human activity in designated locations, such as national parks, nature reserves, or marine sanctuaries, with the aim of maintaining and restoring biodiversity. Whether PAs generally achieve this aim remains unclear. Several broad-scale (i.e., global or continental) studies have investigated the effectiveness of terrestrial and marine PAs, providing insights into their potential for reducing biodiversity loss 5– 7, Received: 27 March 2025 Accepted: 21 November 2025 Check for updates A full list of af ﬁliations appears at the end of the paper. e-mail: [email protected] Nature Communications| (2025) 16:11146 1 1234567890():,; 1234567890():,; exploitation8, and habitat loss 9,10. However, similar broad-scale per- spectives are currently lacking for freshwaters. This knowledge gap is particularly concerning given that freshwater ecosystems harbor a disproportionate amount of global biodiversity by area, and this bio- diversity is declining faster compared to terrestrial and marine ecosystems 11– 13. Evidence of PA effectiveness for individual freshwater ecosystems, or freshwaters in individual regions, is currently mixed14, with some PAs generally bene ﬁtting freshwater biodiversity 15,16 whereas others exhibit little to no effect17– 19. This variability highlights the need for research that evaluate s the general effectiveness of PAs for freshwater biodiversity at broader spatial scales. Inland (i.e., non-marine) PAs may broadly fail to protect fresh- water biodiversity because their boundaries and management typically prioritize terrestrial habitats and charismatic taxa 20,21, lack explicit goals for freshwaters, and neglect the needs of freshwater taxa22– 24.F o r example, most inland PAs are small, with ~85% less than 10 square kilometers25. However, many freshwater ecosystems, particularly lar- ger rivers, can extend across tens to hundreds of kilometers with catchments encompassing thousands of square kilometers 26.T h u s , local river communities can be impacted by upstream terrestrial pol- lutants and other inputs across broad spatial scales 27– 29.S m a l l - s c a l e protection of a river site can therefore be compromised by inputs arriving from upstream, unprotected areas 30,31. Small PAs may also only succeed at protecting local habitat, while other key upstream and downstream habitats used by mobile freshwater taxa remain unprotected 32– 34. Evaluating the bene ﬁts of inland PAs for freshwaters requires appropriate counterfactuals, i.e., unprotected areas, for comparison. Studies often rely on spatial comparisons of protected and unpro- tected sites 17,19,35,36, but this may produce biased results due to spatial biases in PA placement. For example, PAs tend to be designated in less impacted, forested, higher elevation areas with little human development 37,38, which may already have high and/or stable biodi- versity compared to unprotected sites. These biases often cannot be fully controlled, which makes it dif ﬁcult to distinguish the effects of protection from pre-existing differences between sites 39.A na l t e r - native approach is to incorporate a temporal component into the spatial comparisons, speciﬁcally by comparing the rate of biodiversity change between protected and unprotected sites6.T h i sm e t h o db e t t e r evaluates PA effectiveness by using earlier years within sites as the baseline, thus helping determine whether establishing or expanding protection affected biodiversity, and whether biodiversity was lost (or gained) at a faster rate in unprotected sites. However, such temporal comparisons are hindered by the scarcity of high-resolution time- series data. To address the need for broad-scale, temporal evaluations of PA effectiveness for freshwater biodiversity, we examine 1,754 time series of river invertebrate communities collected between 1986 and 2022 across ten European countries (Fig.1 and Supplementary Table 1). We focus on river invertebrates because they are key components of freshwater biodiversity that provide important ecosystem functions and services 40, and they exhibit consistent compositional responses to human pressures 41. These taxa are therefore commonly used as bioindicators and have been monitored globally for decades, including in countries across Europe. Consequently, analysis of European, long- term river invertebrate community data can address the need for broad-scale, temporal evaluations of PA effectiveness for freshwater biodiversity. We ﬁrst quantify biodiversity change as site-speci ﬁc temporal changes in invertebrate abundance, taxonomic richness, and ecologi- cal quality (a measure of human impacts based on similarity to com- munities in least-impacted conditions; see “Methods” section). We then determine whether the rate o f biodiversity change differs between sites with and without upstream PAs (as in ref.6 for terrestrial A protected area is in the catchment

No -Y e s Proportion protected

0–0.2 0.2–0.4 0.4–0.6 0.6–0.8 0.8–1 N Fig. 1 | Locations of the 1754 sampled European river sites.Sites are in Belgium, Czechia, Denmark, Finland, France, Hungary, Lithuania, Spain, Sweden, and the UK. Sites are colored based on the presence of a protected area in the full upstream catchment (no = red; yes = blue). Point sizes for sites with upstream protected areas are based on the proportion of the catchment covered by protected areas. Article https://doi.org/10.1038/s41467-025-67125-5 Nature Communications| (2025) 16:11146 2 and marine ecosystems), under the expectation that protection would better maintain biodiversity and lead to greater increases in biodi- versity through time. To compare the effects of protection close to a river site versus across its broader catchment, we investigate relationships at four progressively larger upstream distances, ranging from PAs up to 1 km upstream (i.e., the immediate vicinity of a site) 10 km, 100 km, and the full upstream catchment. Lastly, for sites with upstream PAs, we determine whether biodiversity change depends on Effect of protection

Protected - Unprotected - Gain - No gain Effect of expanding protection AbundanceRichnessEQR Rate of biodiversity change (% year-1) a d b e c f

** 871 883 1,262 492 1,450 304 1,472 282 400 1,354 676 1,078 882 872 906 848 Fig. 2 | Overall effects of protected areas on river biodiversity.Rate of temporal change ina, d abundance,b, e richness, andc, f ecological quality (as the Ecological Quality Ratio; EQR) in (a– c) protected and unprotected sites, and (d– f)i ns i t e st h a t gained and did not gain upstream PA cover, for the 1-km, 10-km, 100-km, and full upstream scales. Points show the predicted group mean based on the respective linear mixed model, with lines as 95% conﬁdence intervals. Asterisks indicate signiﬁcant differences between groups based on Likelihood Ratio Tests and cor- rected for multiple comparisons using the Benjamini– Hochberg false discovery rate (c,1k m : P <0 . 0 0 1 ,1 0k m :P =0 . 0 2 1 ;e,1k m : P = 0.039). Numbers in (c, f) indicate the number of sites out of 1754 total in each group, and these same sample sizes apply to (a, b)a n d(d, e). Article https://doi.org/10.1038/s41467-025-67125-5 Nature Communications| (2025) 16:11146 3 the amount of PA cover, or the degree of PA gain, and whether it varies with river size and initial ecological quality. Regarding river size, as discussed above, larger rivers integrate inputs across longer distances, thus potentially exposing their communities to cumulative pollutants from rural and urban sources, so we expect that biodiversity in larger rivers primarily responds to PA cover that spans larger upstream scales. Regarding ecological quality, PAs tend to be designated in already less impacted areas (i.e., better initial ecological quality), and we expect that the effectiveness of such PAs differs from those established in poorer quality sites, which generally have lower biodiversity 42 and thus more scope for improvement. Here, we show that upstream PAs primarily bene ﬁt poor-quality communities where PAs encompass a larger proportion of the catch- ment. These communities exhibit much higher rates of biodiversity recovery than likely would have occurred in the absence of protection. In contrast, PAs have little to no effect on biodiversity in moderate- and high-quality communities, although the latter group may have been unaffected because human impacts in such rivers are generally low regardless of protection status. Our results underscore the need to broadly improve PA effectiveness in freshwaters by ensuring PA design and management explicitly consider freshwater biodiversity and integrate the needs of freshwater ecosystems. Results Protected and unprotected sites Protected and unprotected sites only differed in the rate of ecological quality change (represented as the Ecological Quality Ratio; EQR), and only when protections encompassed smaller upstream scales, speci- ﬁcally when PAs were within 1-km (based on a signi ﬁcant Likelihood ratio test or LRT, n =1 7 5 4 ,L = 23.4, df = 1, P < 0.001) and 10-km upstream distances from a river site (LRT, n =1 7 5 4 ,L =5 . 9 7 , d f=1 , P =0 . 0 3 9 ;F i g .2a– c). However, these changes were always greater (i.e., better) in unprotected sites, which was the opposite of our expecta- tions. For example, the rate of EQR change for protected sites at the 1-km upstream scale was +1.1% year −1, whereas it was +1.9% year −1 for unprotected sites (Fig. 2c). For all other metrics and upstream scales, we found no differences in biodiversity change between protected and unprotected sites (Fig. 2a– c; Supplementary Table 2), with similar proportions of sites in these groups both gaining and losing biodi- versity (Supplementary Fig. 1). For sites with upstream PAs, higher PA cover was related to greater increases in taxon richness and ecological quality, but the nature of these relationships depended on initial ecological quality and upstream scale, as evidenced by signiﬁcant PA cover*ecological quality interactions from generalized additive mixed models (Supplementary Table 3), and differences in effect sizes among upstream scales. Rich- ness primarily increased with higher PA cover close to a site (i.e., at smaller upstream scales), and primarily in initially degraded commu- nities (i.e., lower initial ecological quality; Fig. 3). For example, con- sidering richness at the 1-km upstream scale and an initially poor ecological quality of 0.2 (i.e., 20% similarity to reference conditions), increasing PA cover from <1% to 100% almost tripled the rate at which richness increased, from +2.8% year −1 to +8.2% year −1 (Fig. 3a). These effects weakened as the upstream scale and initial ecological quality increased (Fig.3b– d) to the point that, at the full upstream scale and an initially high ecological quality of 0.8, increasing PA cover from <1% to 100% only increased the rate at which richness increased from +0.7% to +1.1% year −1 (Fig. 3d). Similar to richness, ecological quality also increased with higher PA cover primarily in initially poorer quality communities. However, the effect of upstream scale was the oppo- site to that observed for richness, with greater improvements in eco- logical quality when PA cover increased at larger upstream scales Rate of richness change (% year-1) Upstream protected area (%) a - 1 km b - 10 km c - 100 km d - Full Initial EQR 0.2 0.4 0.6 0.8 Fig. 3 | Effects of the amount of protected area on taxon richness.Relationship between increasing the amount of protected area (PA) cover and the rate of tem- poral change in richness for only sites with PAs at thea 1-km, b 10-km,c 100-km, and d full upstream scales. Lines show the best-ﬁt relationships, with shaded areas as 95% conﬁdence intervals, based on generalized additive mixed models. Line and shading color illustrate how relationships depend on initial ecological quality (as the initial Ecological Quality Ratio, EQR) using examples of 0.2 (red), 0.4 (orange), 0.6 (light blue), and 0.8 (dark blue), which respectively indicate higher to lower human impacts. Article https://doi.org/10.1038/s41467-025-67125-5 Nature Communications| (2025) 16:11146 4 (Supplementary Fig. 2), indicating ecological quality primarily responded to the amount of protection across the catchment. Increasing PA cover did not affect the rate of change in abundance at any upstream scale, and we found no evidence for effects of river size (Supplementary Table 3). Sites that gained and did not gain protected areas Similar to the protected and unprotected sites, sites that gained and did not gain PA cover only differed in the rate of ecological quality change, and only at the 1-km upstream scale (LRT, n =1 7 5 4 ,L =7 . 7 9 , df = 1, P = 0.021), with greater increases in sites that did not gain PA cover (Fig. 2d– f). For all other metrics and upstream scales, we found no differences in biodiversity change between sites that gained and did not gain PA cover (Fig. 2d– f; Supplementary Table 2), with similar proportions of sites in these groups both gaining and losing biodi- versity (Supplementary Fig. 1). For sites that gained upstream PAs, higher gains translated to greater increases in richness and ecological quality, primarily in initi- ally poorer quality communities and only at larger upstream scales (richness: full only, Supplementary Fig. 3; EQRs: 10 km, 100 km, and full, Fig. 4). These increases were stronger for ecological quality and weaker for richness. For example, at the full upstream scale and an initially poor ecological quality of 0.2, increasing the rate of PA gain from <1% year −1 to the maximum observed value of 7.5% year −1 more than tripled the rate of EQR gain, from +4.1% year −1 to +14% year −1 (Fig. 4d). The rate of richness gain almost doubled under the same conditions, from +2.3% year−1 to +4.5% year−1 (Supplementary Fig. 3). Furthermore, as initial ecological quality increased, we found some instances where higher PA gains translated to slightly lower rates of increase in both ecological quality and richness. Using the full upstream scale as an example and an initially high ecological quality of 0.8, increasing the rate of PA gain from <1% to 7.5% year −1 decreased the rate of EQR gain from +0.32% to +0.19% year−1 (Fig. 4d). Increasing PA gain did not affect the rate of change in abundance at any upstream scale, and we found no evidence for effects of river size (Supplementary Table 3). Discussion A principal question for evaluating the effectiveness of protection for nature conservation is to determine what would have happened in its absence 39. Our results show that, broadly speaking, the same changes in river invertebrate biodiversity occurred regardless of the presence or degree of upstream protection, although PAs improved biodiversity outcomes in a subset of poor-quality communities that had or gained PA cover across a larger proportion of their upstream catchment. Additionally, some rivers lost invertebrate biodiversity during our 1986 to 2022 study period, which occurred in a comparable proportion of protected and unprotected sites. We therefore found no consistent evidence that inland PAs have generally bene ﬁted European river invertebrate biodiversity, suggesting that PAs may have not beneﬁted water or habitat quality, given that invertebrates are key indicators of both 43.T h e s eﬁndings provide continental-scale support for similar Rate of EQR change (% year-1) Rate of upstream protected area gain (% year-1) No effect of PA gain at this upstream scale a - 1 km b - 10 km c - 100 km d - Full Initial EQR 0.2 0.4 0.6 0.8 Fig. 4 | Effects of the rate of protected area gain on ecological quality. Rela- tionship between gain of upstream protected area (PA) cover and the rate of temporal change in ecological quality (represented as the Ecological Quality Ratio; EQR) for only sites that gained PA cover at thea 1-km, b 10-km, c 100-km, andd full upstream scales. Lines show the best-ﬁt relationships, with shaded areas as the 95% conﬁdence intervals, based on generalized additive mixed models. Line and shading color illustrate how relationships depend on initial ecological quality using example initial EQRs of 0.2 (red), 0.4 (orange), 0.6 (light blue), and 0.8 (dark blue), which respectively indicate higher to lower human impacts. Black lines and grey shading indicate non-signiﬁcant (P > 0.05) relationships based on Wald-type tests. Best-ﬁt relationships are shown up to the maximum rate of PA gain observed at each upstream scale. Article https://doi.org/10.1038/s41467-025-67125-5 Nature Communications| (2025) 16:11146 5 results from individual freshwater ecosystems and speciﬁcr e g i o n sf o r invertebrates18,31, other taxonomic groups (e.g., ﬁsh17– 19), and water quality18. This conclusion should not be misconstrued as suggest- ing that PAs are ineffective, particularly given that it is based on a subset of total freshwater biodiversity and does not address whether PAs achieved the terrestrial conservation goals they are typically designed and managed for, such as reducing habitat loss 9,10. We also found that PAs increased the rate of improvement in biodiversity and ecological quality for some river invertebrate communities, and other studies have shown PAs beneﬁting certain, individual freshwater eco- systems and taxonomic groups 14,44.O u rﬁndings do, however, highlight the need to broadly improve the capacity of inland PAs to support freshwater biodiversity. Our results for poor-quality communities (e.g., around an initial EQR of 0.2 or 20% similarity to reference conditions) suggest that PAs led to greater increases in biodiversity in these sites than would have occurred without protection. The lesser in ﬂuence of protection on higher quality communities potentially reﬂects the already low human impacts in these sites, thus biodiversity remained high and stable regardless of protection status. This explanation ﬁts with our results showing low PA effectiveness in high-quality communities (e.g., initial EQR around 0.8) where biodiversity was likely already high, and may explain why protection was sometimes associated with lower biodi- versity gains, which may occur if PAs are placed in areas with a lower scope for improvement (e.g., remote, forested catchments 37,38). How- ever, it does not explain why PAs were less effective for moderate- quality communities (e.g., initial EQR around 0.4 – 0.6), which have considerable potential for further improvement. A more likely expla- nation for these communities is that current approaches to imple- menting inland PAs, which typically focus on management of terrestrial habitats 23, can address some stressors affecting poor-quality rivers, but not other stressors that may be more relevant in higher quality ecosystems. For example, land-use change and pollution are among the principal stressors driving freshwater biodiversity loss 45. PAs have some capacity to address these stressors by reducing the human activities that cause them, such as deforestation, urban expansion, intensive agriculture, and tourism 10,46. Doing so can sub- sequently improve water and habitat quality in hydrologically con- nected rivers 47,48. However, as communities recover, other unaddressed stressors may become more relevant, such as upstream ﬂow alterations or climate change 49, thereby limiting PA effectiveness. Maximizing the bene ﬁts of PAs for freshwater biodiversity requires that existing management regimes consider both terrestrial- and freshwater-focused actions 23,50,a n ds e ts p e c iﬁc goals to address the most important stressors in each freshwater ecosystem. Preventing degradation, including in higher quality rivers, also requires con- servation actions beyond establishing PAs, such as better wastewater treatment, habitat restoration, and further improvements to land management practices, including reducing micropollutants 11,51. PA bene ﬁts in initially poor-quality communities varied among upstream scales and community metrics, suggesting that the spatial scale of protection determined which community components were affected. Richness primarily responded to the amount of PA cover close to a site (i.e., at smaller upstream scales), whereas ecological quality primarily responded when protection encompassed and expanded across the broader catchment (i.e., at larger upstream scales). Abundance exhibited no response to protection whatsoever. Increases in richness that neither affect abundance nor substantially alter compositional metrics, such as ecological quality, can occur when only numerically rare species increase 52.S i m i l a r l y ,c o m p o s i t i o n a l changes may not affect richness or abundance if new taxa replace previous taxa53. Our results could therefore be explained by protection at smaller scales primarily increasing rare taxa, and protection across larger scales producing stronger compositional recovery by replacing tolerant with sensitive taxa. Increasing rare taxa can provide some beneﬁts, including potentially diversifying and stabilizing ecosystem functions 54,55, but may represent a less desirable outcome compared to substantially improving invertebrate ecological quality, which is a principal indicator of European river health. Therefore, our results suggest that protecting the broader catchment, or at least a large proportion of the catchment and a river ’s lateral buffer zones, may elicit greater biodiversity beneﬁts. This conclusion supports the value of catchment-scale rather than local-scale approaches to freshwater protection 22,56, including PAs that are con ﬁgured to protect and con- nect key longitudinal (upstream to downstream), lateral (riparian and ﬂoodplain), and vertical habitats (surface to groundwater) 14,33,57. An additional solution to improving PA effectiveness for fresh- waters could be to further limit human activities within current PA boundaries, given that many still permit continued human use 58,s u c h as land development and resource extraction. Designating stricter PAs that do not allow such activities may reduce human impacts 46,t h u s potentially beneﬁting downstream freshwaters. However, evidence that the strictness of a PA ’s designation determines its conservation beneﬁts is equivocal59, including in freshwater ecosystems16,44.S t r i c t e r protection can also counterintuitively lead to worse conservation outcomes by disenfranchising local communities and promoting ille- gal use of protected resources 60. Integrating terrestrial with freshwater approaches to PA design and management may be an alternative approach for improving freshwater conservation outcomes 14,50. Freshwater-focused PAs (e.g., Ramsar wetlands or river PAs61)c a nb e designed based on the distribution of both terrestrial and freshwater biodiversity while accounting for habitat connectivity and down- stream impacts22,30,50. Effective, adequately funded, and co-produced management is also key to PA effectiveness14,60. We therefore advocate that freshwater ecosystems would further beneﬁtf r o mi n c l u s i o ni nP A management priorities that integrate the freshwater needs of local communities and stakeholders. Inland PAs are increasing globally, supported by the 30% by 2030 coverage target set by the Kunming-Montreal Global Biodiversity Framework 4. These PAs typically prioritize the needs of terrestrial habitats and taxa, raising questions about their beneﬁts for freshwater biodiversity. Our ﬁndings, based on European river invertebrate communities, show that PAs have bene ﬁted certain freshwater com- munities, speci ﬁcally poor-quality communities where protection encompassed a larger proportion of the upstream catchment. All other communities exhibited more limited (or no) effects of protection, although the lack of effect in high-quality communities may have occurred because these communities are less impacted regardless of whether they are protected or not. Improving overall PA effectiveness, particularly in impacted rivers, requires design and management strategies that explicitly integrate the needs of freshwater ecosystems 14,57, including actions that address multiple stressors and continuous coverage that extends over larger upstream distances and lateral buffer zones. Accordingly, a holistic, catchment-scale frame- work for managing freshwaters is required14,22,23,62.S u c haf r a m e w o r k would better support freshwater biodiversity, including aquatic invertebrates and the ecosystem functions they provide (e.g., prey, nutrient cycling, and decomposition 40), and would bene ﬁt terrestrial ecosystems via aquatic-terrestrial linkages63 and marine ecosystems via freshwater-marine linkages64. Consequently, improving freshwater protection is a critical issue relevant to all ecosystems and is essential to bend the curve of global biodiversity loss. Methods River invertebrate biodiversity We collated river invertebrate time series from ref. 42 and from data provided by European freshwater researchers and managers. We deﬁned the following criteria for data inclusion: (i) time series must span a duration of ≥10 years with ≥7 individual sampling years to enable robust estimation of biodiversity change; (ii) within a time Article https://doi.org/10.1038/s41467-025-67125-5 Nature Communications| (2025) 16:11146 6 series, samples in different years must be collected using the same methods and from the same three-month season; (iii) data were available at the community-level with taxa identi ﬁed to a consistent taxonomic level through time (if inconsistent levels were used then taxa were adjusted to the most temporally consistent level); and (iv) ecological quality values could be calculated for each community fol- lowing methods compliant with the EU Water Framework Directive (see Supplementary Table 4). These criteria allowed the inclusion of data from ten European countries (Fig.1). Included data encompassed 1754 sites and 24,245 individual years collected between 1986 and 2022. Included time series spanned a mean total duration of 19.7 ± 5.7 years (mean ± SD) with 13.8 ± 5.5 sampling years (Supplementary Table 1). Taxonomic resolution varied among sites, with 57% (993 sites) identiﬁed only to the family level or higher, and 43% (761 sites) iden- tiﬁed to a mixed resolution, typically a combination of families, genera, and species, with some classi ﬁed to intermediate (e.g., subfamily) or higher levels (e.g., Oligochaeta at subclass). These taxonomic differ- ences among sites did not in ﬂuence our results (see Supplementary Fig. 4). Identi ﬁcations higher than species level introduce some uncertainty, given that we cannot detect potential species shifts occurring within these groups. However, such identi ﬁcations still reliably re ﬂect overall community responses to environmental change 65,66 and are common in invertebrate research in which many taxa cannot be reliably identiﬁed to the species level. We quantiﬁed biodiversity for each site and year based on three community metrics: (i) abundance (total number of individuals), (ii) richness (total number of taxa), and (iii) ecological quality, quantiﬁed as the Ecological Quality Ratio (EQR). Ecological quality is commonly used in Europe as a community-based indicator of human impacts, particularly organic pollutants and general environmental degradation 67.I tr e ﬂects the compositional similarity of sensitive and tolerant taxa to expected values from least-impacted reference com- munities, which are de ﬁned based on country-speci ﬁcc r i t e r i a( S u p - plementary Table 4). EQRs are a continuous observed-to-expected ratio of this similarity, ranging from 0 (low similarity indicating high human impacts) to 1 (equal to reference conditions indicating low human impacts), although EQRs for some communities can be above 1, reﬂecting conditions better than the average reference state. We chose EQRs over other compositional metrics, such as temporalβ-diversity, because they provide meaningful information not just about whether communities changed, but also how they changed. Rates of temporal change in each community metric were quan- tiﬁed for each site by relating site-speci ﬁc abundance, richness, and EQRs to sampling year using thegls function from thenlme package in R 68,69, then extracting the slope of this relationship. We included aﬁrst- order autoregressive structure in each model to control for temporal autocorrelation between successive years. All slopes were converted to percent change per year by log-transforming all metrics prior to modeling, then exponentiating the slopes, subtracting 1, and multi- plying by 100. This transformation ensured all rates of biodiversity change had the same units across sites and metrics. Protected areas and upstream scales We obtained vectorial cartographic polygons for inland PAs from Protected Planet 25. We excluded 2% of European PAs (accounting for 6% of total cover) for which the year of establishment was unknown. We further excluded all point data due to analytical errors that arise when inferring the dimensions of PAs with unknown boundaries 70.T h e majority of point data in our included countries (1171 points out of 1247 total or 94%) were natural monuments in Sweden. These PAs have a registered area of 0 km 2 because they are individual features, such as a single tree or rock formation, thus contributing marginally to total protected area cover. Of the remaining 76 excluded points, 20 were Ramsar wetlands with a total area of 296 km 2, and 50 were large bio- sphere reserves with a total area of 94,188 km 2.T o ﬁll the biosphere information gap, we used data on European biosphere boundaries from ref. 71, although we excluded the outer transition zones, which are not considered protected. Polygons for all included PAs were dis- solved into a single layer, with no distinctions made between different PA types (discussed further in Supplementary Note 1). In addition to the PA polygons, for each site, we produced upstream polygons representing four different spatial scales. The four scales included: lateral buffer zones extending up the main channel and all tributaries to (i) 1-, (ii) 10-, and (iii) 100-km longitudinal dis- tances; and (iv) the full upstream contributing area, i.e., the upstream catchment including all terrestrial areas that drain into a site. Each upstream scale was selected to account for the potential effects of PAs at progressively greater distances from each site, with the full upstream scale also accounting for PA effects outside the lateral buffer zones. Upstream distances were delineated using the Hydro- graphy90m river network and the hydrographr package 72,73.B u f f e r zone widths for 1 – 100-km upstream scales were quantiﬁed as 100 m multiplied by the Strahler order of each segment, plus half the pre- dicted river width based on its order from ref.74 (see Fig. 5). We used this approach because human activities adjacent to a river typically exert the strongest inﬂuence on local communities, and 100-m lateral buffer zones effectively capture these effects 75– 77. Additionally, our data encompassed sites from rivers of Strahler orders 1 – 11, repre- senting small streams to very large rivers, and larger lateral areas are needed to capture the larger surface and ground water inputs to higher-order rivers 78. Lastly, including river width as part of the buffer zones captured PAs that encompass rivers. The full upstream con- tributing area was delineated using the get_upstream_catchmentfunc- tion from the hydrographrpackage. Based on the PA and upstream polygons, we calculated the per- cent of each upstream scale covered by PAs to represent both the presence (>0% cover) and degree (total % cover) of protection. We also calculated the rate of temporal change in percent PA cover to capture biodiversity responses to PA expansion. PA cover was calculated for the year before the ﬁrst and last year of each invertebrate time series, Fig. 5 | Illustrated example of the four upstream scales. A river site (red circle) and the upstream scales captured by lateral buffer zones (dashed lines) extending up to 1-km (purple), 10-km (pink), and 100-km (yellow) upstream distances. The solid black outline represents the full upstream contributing area. Note the decline in buffer zone thickness from higher (larger) to lower (smaller) order rivers. Article https://doi.org/10.1038/s41467-025-67125-5 Nature Communications| (2025) 16:11146 7 which allowed invertebrate communities≥1 year to respond to envir- onmental changes resulting from PA establishment. Percent PA cover was quantiﬁed as the mean percent cover between the ﬁrst and last years (always ranging from 0 to 100% across sites). Temporal changes in percent PA cover (% year −1) for each site were quantiﬁed as the slope of the relationship between PA cover and year, which ranged from 0 to 9% year −1 depending on the upstream scale (1 km: 0 – 9%; 10 km: 0– 8.6%; 100 km: 0– 7.5%; Full: 0– 7.5% year−1). PA cover changes were only neutral or positive because Protected Planet data cannot repre- sent PA cover declines 79, although PA cover has generally increased globally58. In addition to PA cover, we quantiﬁe dt h es i z e( i nk m2)o fe a c hf u l l upstream area to represent river size, given that larger rivers have larger upstream areas. Size was calculated based on the number of 90 m by 90 m pixels in the full upstream area, derived from the Hydrography90m river network. Sites primarily encompassed med- ium- to larger-sized rivers, with 671 sites out of 1754 total having an upstream catchment size between 10 and 100 km 2 and 701 sites between 100 and 1000 km 2, with the remainder comprised of very small (135 sites <10 km 2) and very large rivers (247 sites

1000 km2). Statistical analyses We conducted two sets of analyses: (1) categorical comparisons of changes in abundance, richness, and EQRs between protected and unprotected sites (i.e., those with and without upstream PAs), and between sites that did or did not gain upstream PA cover; and (2) for sites with upstream PAs, we related the rate of biodiversity change to the amount of upstream PA cover and the rate of PA gain using regression. The ﬁrst set of analyses provided a broad overview of the effects of having or gaining any upstream PA cover. The second set determined whether the degree of protection, or its rate of gain, inﬂuenced biodiversity change. We also used the second set of ana- lyses to test the in ﬂuence of river size and initial ecological quality (detailed below). These temporal trend comparisons have some strengths compared to other potential approaches, such as before- and-after comparisons or spatial comparisons of protected and environmentally similar unprotected sites. Speciﬁcally, using temporal trends of percent biodiversity change enables comparison of sites that differ in total biodiversity, and allows for variation in protection timing (e.g., sites can be already protected at the start of their time series or can become protected later). Temporal analyses also allow for changes in protection effectiveness through time, such as lagged effects, and capture the potential compounding effects of establishing multiple PAs in subsequent years. For our ﬁrst set of analyses, we compared sites with upstream PAs (>0% cover; ‘protected’) versus those without (0% cover; ‘unprotected’), and those that gained upstream PAs (>0% cover year −1; ‘gain’) versus those that did not (0% cover year −1; ‘no gain ’). Sites were assigned to these categories separately for each upstream scale, given that sites could change assignments across scales (e.g., an upstream PA is present within 10 km but not 1 km). We then related the continuous, site-level rates of biodiversity change to these categories using linear mixed models (LMMs) conducted in the lme4 package

Additionally, each model included ﬁxed continuous terms for site latitude and longitude to control for broad-scale spatial trends, a ﬁxed continuous term for time-series length to control for slope differences among shorter to longer time series, and a random intercept term designating the provider of each dataset to control for differences in sampling methods among providers (see Supplementary Table 1). We tested the signi ﬁcance ( P < 0.05) of the ﬁxed categorical PA term by dropping it from each model and comparing the reduced versus fuller models using Likelihood ratio tests 81.W e ran separate models for each upstream scale. To control for conducting multiple models using the same response variables, we corrected all P-values using the Benjamini– Hochberg false discovery rate82. For our second set of analyses, we used generalized additive mixed models (GAMMs) to relate biodiversity change to the amount of upstream PA cover for sites with upstream PAs, and to the rate of PA gain for sites that gained upstream PAs. Models were conducted in the mgcv package
PA cover and rates of gain were converted to proportions and square-root transformed prior to analyses to pro- duce a more even distribution of values. To determine the in ﬂuence of river size, we included an interaction between the PA term and a continuous term for the size of the full upstream area (log 10-trans- formed km 2). To determine the in ﬂuence of initial ecological qual- ity, we included an interaction between the PA term and a continuous term for the EQR averaged across the ﬁrst three sam- pling years to represent the initial status of the community. The individual PA, river size, and quality terms were modeled as ﬁxed smoothed terms using thin-plate regression splines, with all ﬁxed interactions modeled using tensor product smooths. Additionally, we included the same ﬁxed and random control variables as for the LMMs. All smoothed terms used the default basis dimensions, and we checked model diagnostics, including the need for higher basis dimensions, using the gam.check function. All GAMMs used a Gaussian distribution (identity link function). The signi ﬁcance (P < 0.05) of interactions, and if needed the individual PA term, were determined using Wald-type tests
We corrected all P-values as above when conducting multiple models using the same response variables. Reporting summary Further information on research design is available in the Nature Portfolio Reporting Summary linked to this article. Data availability All data needed to repeat our analyses are publicly available from https://doi.org/10.6084/m9.ﬁgshare.25245430. Code availability All code needed to repeat our analyses is publicly available from https://doi.org/10.6084/m9.ﬁgshare.25245430. References
Ceballos, G. et al. Accelerated modern human –induced species losses: entering the sixth mass extinction.Sci. Adv. 1, e1400253 (2015).
Cowie, R. H., Bouchet, P. & Fontaine, B. The Sixth Mass Extinction: fact, ﬁction or speculation?.Biol. Rev. 97, 640–663 (2022).
Sayer, C. A. et al. One-quarter of freshwater fauna threatened with extinction.Nature 638,1 3 8–145 (2025).
CBD (Convention on Biological Diversity). The Kunming-Montreal Global Biodiversity Framework. https://www.cbd.int/doc/ decisions/cop-15/cop-15-dec-04-en.pdf(2022).
Geldmann, J. et al. A global analysis of management capacity and ecological outcomes in terrestrial protected areas.Conserv. Lett.11, e12434 (2018).
Langhammer, P. F. et al. The positive impact of conservation action. Science 384,4 5 3–458 (2024).
Gray, C. L. et al. Local biodiversity is higher inside than outside terrestrial protected areas worldwide.Nat. Commun. 7,1 2 3 0 6 (2016).
Sciberras, M. et al. Evaluating the relative conservation value of fully and partially protected marine areas. Fish. Fish. 16, 58–77 (2015).
Li, G. et al. Mixed effectiveness of global protected areas in resisting habitat loss. Nat. Commun. 15, 8389 (2024). Article https://doi.org/10.1038/s41467-025-67125-5 Nature Communications| (2025) 16:11146 8
Leberger, R., Rosa, I. M. D., Guerra, C. A., Wolf, F. & Pereira, H. M. Global patterns of forest loss across IUCN categories of protected areas. Biol. Conserv. 241,1 0 8 2 9 9( 2 0 2 0 ) .
Tickner, D. et al. Bending the curve of global freshwater biodiversity loss: an emergency recovery plan.BioScience 70, 330–342 (2020).
Albert, J. S. et al. Scientists ’warning to humanity on the freshwater biodiversity crisis.Ambio 50,8 5–94 (2021).
Dudgeon, D. & Strayer, D. L. Bending the curve of global freshwater biodiversity loss: what are the prospects?.Biol. Rev. 100,2 0 5–226 (2024). 1 4 . A c r e m a n ,M . ,H u g h e s ,K .A . ,A r t h i n g t o n ,A .H . ,T i c k n e r ,D .&D u e ñ a s , M.-A. Protected areas and freshwater biodiversity: a novel sys- tematic review distils eight lessons for effective conservation. Conserv. Lett. 13, e12684 (2020).
Koning, A. A., Perales, K. M., Fluet-Chouinard, E. & McIntyre, P. B. A network of grassroots reserves protects tropical riverﬁsh diversity. Nature 588,6 3 1–635 (2020). 1 6 . V a l e n t i m ,H .I .L . ,F e i o ,M .J .&A l m e i d a ,S .F .P .A s s e s s i n gt h e effectiveness of terrestrial protected areas towards riverine eco- systems. Sci. Total Environ. 958,1 7 7 8 7 8( 2 0 2 5 ) .
Chessman, B. C. Do protected areas bene ﬁt freshwater species? A broad-scale assessment forﬁsh in Australia’sM u r r a y–Darling Basin. J. Appl. Ecol. 50,9 6 9–976 (2013).
Nogueira, J. G., Teixeira, A., Varandas, S., Lopes-Lima, M. & Sousa, R. Assessment of a terrestrial protected area for the conservation of freshwater biodiversity.Aquat. Conserv. Mar. Freshw. Ecosyst.31, 520–530 (2021).
Gavioli, A., Filipe, A. F., Patonai, K., Milardi, M. & Castaldelli, G. Effectiveness of the Natura 2000 network for freshwaterﬁsh c o n s e r v a t i o ni naM e d i t e r r a n e a nr e g i o n .Front. Environ. Sci. 11, 251 (2023). 2 0 . H e r b e r t ,M .E . ,M c i n t y r e ,P .B . ,D o r a n ,P .J . ,A l l a n ,J .D .&A b e l l ,R . Terrestrial reserve networks do not adequately represent aquatic ecosystems.Conserv. Biol. 24,1 0 0 2–1011 (2010).
Delso, Á, Fajardo, J. & Muñoz, J. Protected area networks do not represent unseen biodiversity.Sci. Rep. 11, 12275 (2021).
Abell, R., Allan, J. D. & Lehner, B. Unlocking the potential of protected areas for freshwaters. Biol. Conserv. 134,4 8 –63 (2007).
Juffe-Bignoli, D. et al. Achieving Aichi Biodiversity Target 11 to improve the performance of protected areas and conserve fresh- water biodiversity.Aquat. Conserv. Mar. Freshw. Ecosyst.26, 133–151 (2016).
Gonçalves, D. V. & Hermoso, V. Global goals overlook freshwater conservation.Science 377,3 8 0–380 (2022).
Protected Planet. Protected planet: the world database on pro- tected areas (WDPA) and world database on other effective area- based conservation measures (WD-OECM).www.protectedplanet. net (2023).
Burgers, H. E., Schipper, A. M. & Jan Hendriks, A. Size relationships of water discharge in rivers: scaling of discharge with catchment area, main-stem length and precipitation.Hydrol. Process.28, 5769–5775 (2014).
Lorenz, A. W. & Feld, C. K. Upstream river morphology and riparian land use overrule local restoration effects on ecological status assessment.Hydrobiologia704,4 8 9–501 (2013). 2 8 . S t o l l ,S . ,B r e y e r ,P . ,T o n k i n ,J .D . ,F r ü h ,D .&H a a s e ,P .S c a l e - dependent effects of river habitat quality on benthic invertebrate communities— implications for stream restoration practice.Sci. Total Environ. 553,4 9 5–503 (2016). 2 9 . Z h a n g ,W . ,S w a n e y ,D .P . ,H o n g ,B .&H o w a r t h ,R .W .A n t h r o p o g e n i c phosphorus inputs to a river basin and their impacts on phosphorus ﬂuxes along its upstream-downstream continuum.J. Geophys. Res.: Biogeosciences122,3 2 7 3–3287 (2017).
Pringle, C. M. Hydrologic connectivity and the management of biological reserves: a global perspective.Ecol. Appl. 11,9 8 1–998 (2001).
Mancini, L. et al. Biological quality of running waters in protected areas: the inﬂuence of size and land use. Biodivers. Conserv.14, 351–364 (2005). 3 2 . D a r w a l l ,W .R .T .e ta l .I m p l i c a t ions of bias in conservation research and investment for freshwater species.Conserv. Lett. 4,4 7 4–482 (2011).
Hermoso, V., Filipe, A. F., Segurado, P. & Beja, P. Filling gaps in a large reserve network to address freshwater conservation needs.J. Environ. Manag. 161,3 5 8–365 (2015).
Grantham, T. E. et al. Missing the boat on freshwater ﬁsh con- s e r v a t i o ni nC a l i f o r n i a .Conserv. Lett. 10 ,7 7–85 (2017).
Coetzee, B. W. T., Gaston, K. J. & Chown, S. L. Local scale comparisons of biodiversity as a test for global protected area ecological performance: a meta-analysis. PLoS ONE 9, e105824 (2014).
Cazalis, V., Belghali, S. & Rodr igues, A. S. L. Using a large-scale biodiversity monitoring dataset to test the effectiveness of pro- tected areas at conserving North-American breeding birds.Peer Community J. 1, e3 (2021).
Joppa, L. N. & Pfaff, A. High and far: biases in the location of pro- tected areas. PLoS ONE 4, e8273 (2009).
Venter, O. et al. Bias in protected-area location and its effects on long-term aspirations of biodiversity conventions.Conserv. Biol.32, 127–134 (2018).
Rodrigues, A. S. L. & Cazalis, V. The multifaceted challenge of evaluating protected area effectiveness.Nat. Commun. 11, 5147 (2020). 4 0 . M a c a d a m ,C .R .&S t o c k a n ,J .A .M o r et h a nj u s tﬁsh food: ecosystem services provided by freshwater insects.Ecol. Entomol. 40, 113–123 (2015). 4 1 . B o n a d a ,N . ,P r a t ,N . ,R e s h ,V .H .&S t a t z n e r ,B .D e v e l o p m e n t si n aquatic insect biomonitoring: acomparative analysis of recent approaches.Annu. Rev. Entomol. 51, 495–523 (2006).
Sinclair, J. S. et al. Multi-decadal improvements in the ecological quality of European rivers are not consistently reﬂected in biodi- versity metrics.Nat. Ecol. Evol. 8,4 3 0–441 (2024).
Carter, J. L., Resh, V. H. & Hanna ford, M. J. Macroinvertebrates as biotic indicators of environmental quality. InMethods in Stream Ecology (eds. Lamberti, G. A. & Hauer, F. R.) 293–318 (Academic Press, The Netherlands, 2017).
Santangeli, A. et al. Mixed effects of a national protected area network on terrestrial and freshwater biodiversity.Nat. Commun. 14, 5426 (2023).
Jaureguiberry, P. et al. The direc t drivers of recent global anthro- pogenic biodiversity loss.Sci. Adv. 8, eabm9982 (2022).
Jones, K. R. et al. One-third of global protected land is under intense human pressure. Science 360,7 8 8–791 (2018).
Harrison, I. J. et al. Protected areas and freshwater provisioning: a global assessment of freshwater provision, threats and manage- ment strategies to support human water security.Aquat. Conserv. Mar. Freshw. Ecosyst.26,1 0 3–120 (2016).
Staponites, L. R., Simon, O. P., Barták, V. & Bílý, M. Management effectiveness in a freshwater protected area: long-term water quality response to catchment-scale land use changes.Ecol. Indic. 144, 109438 (2022).
Haase, P. et al. The recovery of European freshwater biodiversity has come to a halt. Nature 620,5 8 2–588 (2023).
Leal, C. G. et al. Integrated terre strial-freshwater planning doubles conservation of tropical aquatic species.Science 370,1 1 7–121 (2020).
Wolfram, J., Stehle, S., Bub, S., Petschick, L. L. & Schulz, R. Water quality and ecological risks in European surface waters– Article https://doi.org/10.1038/s41467-025-67125-5 Nature Communications| (2025) 16:11146 9 monitoring improves while water quality decreases.Environ. Int. 152, 106479 (2021).
Hurlbert, S. H. The nonconcept of species diversity: a critique and alternative parameters.Ecology 52,5 7 7–586 (1971).
Hillebrand, H. et al. Biodiversity change is uncoupled from species richness trends: consequences for conservation and monitoring.J. Appl. Ecol. 55,1 6 9–184 (2018).
Violle, C. et al. Functional rarity: the ecology of outliers.Trends Ecol. Evol. 32,3 5 6–367 (2017).
Dee, L. E. et al. When do ecosystem services depend on rare spe- cies?. Trends Ecol. Evol. 34,7 4 6–758 (2019).
Cid, N. et al. From meta-system theory to the sustainable man- agement of rivers in the Anthropocene.Front. Ecol. Environ.20, 49–57 (2022).
Nel, J. L., Reyers, B., Roux, D. J. & Cowling, R. M. Expanding pro- tected areas beyond their terrestrial comfort zone: identifying spatial options for river conservation.Biol. Conserv.142,1 6 0 5–1616 (2009).
Guan, Z., Elleason, M., Goodale, E. & Mammides, C. Global patterns and potential drivers of human settlements within protected areas. Environ. Res. Lett. 16, 064085 (2021).
Elleason, M. et al. Strictly protected areas are not necessarily more effective than areas in which multiple human uses are permitted. Ambio 50,1 0 5 8–1073 (2021).
Geldmann, J., Manica, A., Burgess, N. D., Coad, L. & Balmford, A. A global-level assessment of the effectiveness of protected areas at resisting anthropogenic pressures.Proc. Natl. Acad. Sci. USA 116, 23209–23215 (2019). 6 1 . C a ñ e d o - A r g ü e l l e s ,M . ,H e r m o s o ,V . ,H e r r e r a - G r a o ,T . ,B a r q u í n ,J .& Bonada, N. Freshwater conservation planning informed and vali- dated by public participation: the Ebro catchment, Spain, as a case study. Aquat. Conserv. Mar. Freshw. Ecosyst.29,1 2 5 3–1267 (2019).
Haubrock, P. J. et al. A holistic catchment-scale framework to guide ﬂood and drought mitigation towards improved biodiversity con- servation and human wellbeing.WIREs Water 12,e 7 0 0 0 1( 2 0 2 5 ) .
Schindler, D. E. & Smits, A. P. Subsidies of aquatic resources in terrestrial ecosystems.Ecosystems20,7 8–93 (2017).
Broadley, A., Stewart-Koster, B., Burford, M. A. & Brown, C. J. A global review of the critical link between river ﬂows and pro- ductivity in marine ﬁsheries. Rev. Fish. Biol. Fish. 32 ,8 0 5–825 (2022).
Heino, J. & Soininen, J. Are higher taxa adequate surrogates for species-level assemblage patterns and species richness in stream organisms?.Biol. Conserv. 137,7 8–89 (2007).
Qu, Y. et al. Signi ﬁcant improvement in freshwater invertebrate biodiversity in all types of English rivers over the past 30 years.Sci. Total Environ. 905,1 6 7 1 4 4( 2 0 2 3 ) .
Birk, S. et al. Three hundred ways to assess Europe’s surface waters: an almost complete overview of biological methods to implement the Water Framework Directive.Ecol. Indic. 18,3 1–41 (2012).
Pinheiro, J., Bates, D. & R Cor e Team. nlme: Linear and nonlinear mixed effects models https://CRAN.R-project.org/package= nlme (2023).
R. Core Team. R: A Language and Environment for Statistical Com- puting https://www.r-project.org/(R Foundation for Statistical Computing, 2023).
Visconti, P. et al. Effects of errors and gaps in spatial data sets on assessment of conservation progress.Conserv. Biol.27,1 0 0 0–1010 (2013).
Palliwoda, J., Büermann, A., Fischer, J., Kraemer, R. & Schröter, M. Zoning of UNESCO biosphere reserves: a comprehensive set of geodata for Europe. Front. Ecology Evol. 9,6 7 3( 2 0 2 1 ) .
Amatulli, G. et al. Hydrography90m: a new high-resolution global hydrographic dataset.Earth Syst. Sci. Data 14,4 5 2 5–4550 (2022).
Schürz, M. et al. hydrographr: Scalable hydrographic data proces- sing in R. https://glowabio.github.io/hydrographr/(2023).
Downing, J. A. et al. Global abundance and size distribution of streams and rivers. Inland Waters 2, 229–236 (2012).
Sliva, L. & Williams, D. D. Bu ffer zone versus whole catchment approaches to studying land use impact on river water quality. Water Res. 35,3 4 6 2–3472 (2001).
Li, S., Gu, S., Tan, X. & Zhang, Q. Water quality in the upper Han River basin, China: the impacts of land use/land cover in riparian buffer zone. J. Hazard. Mater. 165,3 1 7–324 (2009). 7 7 . N a v a - L ó p e z ,M .Z . ,D i e m o n t ,S .A .W . ,H a l l ,M .&Á v i l a - A k e r b e r g ,V . Riparian buffer zone and whole watershed inﬂuences on river water quality: implications for ecosystem services near megacities. Environ. Process.3,2 7 7– 305 (2016).
Nardi, F., Vivoni, E. R. & Grimaldi, S. Investigating a ﬂoodplain scaling relation using a hydrogeomorphic delineation method. Water Resour. Res. 42, W09409 (2006).
Lewis, E. et al. Dynamics in the global protected-area estate since
Conserv. Biol. 33,5 7 0–579 (2019).
Bates, D., Mächler, M., Bolker, B. & Walker, S. Fitting linear mixed- effects models using lme4.J. Stat. Softw. 67,1 –48 (2015). 8 1 . Z u u r ,A . ,I e n o ,E .N . ,W a l k e r ,N . ,S a v e l i e v ,A .A .&S m i t h ,G .M .Mixed Effects Models and Extensions in Ecology with R574 (Springer, New York, 2009). 8 2 . B e n j a m i n i ,Y .&H o c h b e r g ,Y .C o n t r o l l i n gt h ef a l s ed i s c o v e r yr a t e :a practical and powerful approach to multiple testing.J. R. Stat. Soc. Ser. B Methodol. 57,2 8 9–300 (1995).
Wood, S. N. Generalized Additive Models: An Introduction with R (Chapman and Hall/CRC, New York, 2017). Acknowledgements This study was facilitated by the Senckenberg Synthesis Center on Freshwater Biodiversity Change in Europe and its data providers. We thank Marie-Josée Létourneau for the artwork in Fig.5 and for permitting i t su s e .J . S . S .a n dP . H .r e c e i v e dfunding from the EU Horizon 2020 project eLTER PLUS (#871128). P.H. received funding from the Frontiers Research Foundation (Frontiers Planet Prize). N.J.B. received funding from the Research Council of Lithuania (#S-MIP-24-61) and acknowl- edges the Lithuanian EPA. D.C.-M. received funding from the EU Horizon 2020 research and innovation programme (#10106238). S.D. received funding from the Leibniz Competition (#J45/2018). A.M.G.-F. acknowl- edges the grant RYC2023-045780-I funded by MICIU/AEI/10.13039/ 501100011033 and by ESF+. L.K. received funding from Formas (#2023-

and sourced data from the Swedish University of Agricultural Sciences database (Miljödata). A.L. received funding from the Spanish Ministry of Science and Innovation (#PID2020-115830GB-100). P.P. was funded by the Czech Science Foundation (#GA23-05268S), and thanks to CHMI and the Povodí enterprises for the provided data. Author contributions J.S.S. and P.H. conceived the study. J.A., S.D., and R.B.S. contributed to planning and methods. J.S.S. and P.H. wrote most of the manuscript with contributions from all authors, particularly R.S. and E.A.R.W. J.A., N.J.B., M.C.-A., Z.C., D.C.-M,. M.Fe., M.Fl., M.A.E.F., P.L.M.G., A.M.G.-F., K.-L.H., R.K.J., L.K., A.L., T.M., R.P., P.P., J.J.R., R.V., G.V., and M.W. provided invertebrate data or contributed to calculating ecological quality values for their respective countries. Funding Open Access funding enabled and organized by Projekt DEAL. Competing interests Since April 16th, 2025, Miguel Cañedo-Argüelles has been seconded to the ERC Executive agency. The views expressed in this paper are purely those of the author. They do not necessarily reﬂect the views or ofﬁcial Article https://doi.org/10.1038/s41467-025-67125-5 Nature Communications| (2025) 16:11146 10 positions of the European Commission, the ERC Executive Agency, or the ERC Scientiﬁc Council. The other authors declare no competing interests. Additional information Supplementary information The online version contains supplementary material available at https://doi.org/10.1038/s41467-025-67125-5. Correspondence and requests for materials should be addressed to James S. Sinclair. Peer review information Nature Communications thanks Steve Ormerod and the other anonymous reviewer(s) for their contribution to the peer review of this work. A peer review ﬁle is available. Reprints and permissions information is available at http://www.nature.com/reprints Publisher’s note Springer Nature remains neutral with regard to jur- isdictional claims in published maps and institutional afﬁliations. Open Access This article is licensed under a Creative Commons Attribution 4.0 International License, which permits use, sharing, adaptation, distribution and reproduction in any medium or format, as long as you give appropriate credit to the original author(s) and the source, provide a link to the Creative Commons licence, and indicate if changes were made. The images or other third party material in this article are included in the article’s Creative Commons licence, unless indicated otherwise in a credit line to the material. If material is not included in the article’s Creative Commons licence and your intended use is not permitted by statutory regulation or exceeds the permitted use, you will need to obtain permission directly from the copyright holder. To view a copy of this licence, visit http://creativecommons.org/ licenses/by/4.0/. © The Author(s) 2025 1Department of River Ecology and Conservation, Senckenberg Research Institute and Natural History Museum Frankfurt, Gelnhausen, Germany. 2School of Science and Technology, Nottingham Trent University, Nottingham, UK. 3Division of Biology, Kansas State University, Manhattan, KS, USA. 4Freshwater and Marine Solutions, Finnish Environment Institute, Oulu, Finland. 5State Scientiﬁc Research Institute Nature Research Centre, Vilnius, Lithuania. 6SHE2 Research Group, FEHM-Lab (Freshwater Ecology, Hydrology and Management), Institute of Environmental Assessment and Water Research (IDAEA), CSIC, Barcelona, Spain. 7Department of Hydrobiology, University of Pécs, Pécs, Hungary. 8HUN-REN Balaton Limnological Research Institute, Tihany, Hungary. 9Institute of Aquatic Ecology, HUN-REN Centre for Ecological Research, Budapest, Hungary. 10IFREMER–DYNECO/LEBCO, Centre de Bretagne, Plouzané, France. 11Leibniz Institute of Freshwater Ecology and Inland Fisheries (IGB), Berlin, Germany. 12INRAE, UR RiverLy, centre de Lyon-Villeurbanne, Villeurbanne, Cedex, France. 13University of Paris-Saclay, INRAE, UR HYCAR, Antony, France. 14Department of Animal Sciences and Aquatic Ecology, Ghent University, Ghent, Belgium. 15IHCantabria - Instituto de Hidráulica Ambiental de la Universidad de Cantabria, Santander, Spain. 16Nature Solutions, Finnish Environment Institute, Oulu, Finland. 17Ecology and Genetics Research Unit, University of Oulu, Oulu, Finland. 18Department of Aquatic Sciences and Assessment, Swedish University of Agricultural Sciences, Uppsala, Sweden. 19Department of Forest Ecology and Management, Swedish University of Agricultural Sciences, Umeå, Sweden. 20Department of Plant Biology and Ecology, University of the Basque Country (UPV/EHU), Leioa, Bilbao, Spain. 21Oulanka Research Station, University of Oulu Infrastructure Platform, Kuusamo, Finland. 22Water, Energy and Environmental Engineering Research Unit, Faculty of Technology, University of Oulu, Oulu, Finland. 23Department of Botany and Zoology, Faculty of Science, Masaryk University, Brno, Czech Republic. 24Department of Ecoscience, Aarhus University, Aarhus, Denmark. 25Research Center One Health Ruhr, University Alliance Ruhr & Faculty for Biology, University of Duisburg-Essen, Essen, Germany. 26Flanders Environment Agency (VMM), Aalst, Belgium. 27Department of Tisza Research, Institute of Aquatic Ecology, HUN-REN Centre for Ecological Research, Debrecen, Hungary. 28School of Life Sciences, University of Essex, Colchester, UK. 29Faculty of Biology, University of Duisburg-Essen, Essen, Germany. e-mail: [email protected] Article https://doi.org/10.1038/s41467-025-67125-5 Nature Communications| (2025) 16:11146 11