Declining Streamflow Induces Collapse and Replacement of Native Fish In

RESEARCH COMMUNICATIONS RESEARCH COMMUNICATIONS Declining streamflow induces collapse and 465 replacement of native fish in the American Southwest

Albert Ruhí1*, Julian D Olden2, and John L Sabo1,3

Water scarcity is a global threat to freshwater biodiversity, but connecting variation in streamflow to viability of imperiled faunas remains a challenge. Here we combined time-series modeling techniques on long-term ecohydrological data to quantify flow–ecology relationships on native and non-native riverine fish in the American Southwest, and simulate likely fish trajectories and “quasi-extinction” risks in the near future. Streamflow has been declining conspicuously over the past 30 years in the Colorado and Rio Grande river basins, and year-to- year variation in streamflow influences the covariation between native and non-native fish abundance. Risks of decline are high (>50%) for nearly three-quarters of the modeled native species, and current trends in streamflow increase quasi-extinction risk for natives (+8.5%) but reduce this risk for non-natives (–5.9%). Hydrological changes need to be mitigated if we are to slow down the rapid replacement of native biodiversity with non-native species in American Southwest rivers.

Front Ecol Environ 2016; 14(9): 465–472, doi:10.1002/fee.1424

limate change is projected to increase aridity across spread within already occupied areas (Bernardo et al. C already parched landscapes of the southwestern US 2003; Propst et al. 2008; Mims and Olden 2013). Many over the coming decades (Seager et al. 2013). Precipitation rivers in the American Southwest have witnessed decreas- reductions and higher temperatures, combined with ing streamflow and anomalous timing and magnitude of water deficits that arise from human land use, dams, and low-flow events (Sabo 2014; Ruhí et al. 2015), yet frame- water storage in reservoirs, contribute to diminished flows works for forecasting how climate and hydrological in rivers (Lake 2011). As a result of increased aridity, flow change may affect the future persistence of native and regulation, and over-allocation of freshwater resources, non-native riverine biodiversity are limited. This is a crit- reduced availability of surface water has become the “new ical knowledge gap, given the growing importance of normal” (Sabo et al. 2010). These mounting pressures are these predictions for biodiversity conservation (IPCC manifested in many Southwest rivers, as evidenced by 2014) and the need to quantify the effects of climate decreasing streamflow and increasing frequencies and change and other stressors on biodiversity (McClure et al. magnitudes of low flows, including zero-flow events 2013; Staudt et al. 2013). (Jaeger et al. 2014; Sabo 2014). Growing demands and Our work addresses this knowledge gap by projecting climate-change-induced flow alteration may force even the resilience of fish communities to hydrologic drought. large dryland river systems into permanent hydrological First, we implement a statistical modeling framework to drought (Dettinger et al. 2015). quantify the effects of past hydrological change on native Water scarcity threatens freshwater biodiversity glob- and non-native riverine fish. This community viability ally (Vörösmarty et al. 2010) through changes in water model is transferable to other habitats and facets of biodi- quality and quantity, such as habitat loss and warming versity. Next, we estimate hydrological change and risk of water temperatures, and by decreasing longitudinal habi- quasi-extinction (defined below) for fish across a broad tat connectivity via increased channel drying (Lake region of the southwestern US (the Colorado, Rio 2003; Jaeger et al. 2014; Perkin et al. 2015). Life-history Grande, and Mimbres river basins), based on multi- strategies of aquatic species have evolved with natural decadal time-series datasets. Finally, we extend our flow regimes (Bunn and Arthington 2002; Lytle and Poff framework to forecast near-term viability under two 2004), and evidence shows that altered flow regimes can scenarios (continued versus halted drought conditions), compromise the long-term persistence of native faunas which facilitates determining how hydrological change while concurrently offering opportunities to non-native may differentially affect native and non-native fish. species to invade previously unoccupied areas and to Collectively, our study allows for an improved understanding of linkages between water scarcity and risk of 1Julie Ann Wrigley Global Institute of Sustainability, Arizona biodiversity loss in dryland rivers of the southwestern US. State University, Tempe, AZ *([email protected]); 2School of Such insight may help in guiding water management Aquatic and Fishery Sciences, University of Washington, Seattle, strategies aimed at maximizing native (and minimizing WA; 3School of Life Sciences, Arizona State University, Tempe, AZ non-native) fish resilience to hydrological change.

466 JJ Methods direction to which streamflow in a particular year departed from the characteristic hydrograph. We selected sites with time series of fish community We also asked whether streamflow trends observed over data and historical streamflow data, examined the stream- the target river reaches were representative of the broader flow data to identify historical streamflow anomalies patterns of hydrological change in the American (ie departures from the historical or “normal” hydrograph Southwest. To this end, we selected all stations included [discharge variation over time]) and conducted a large- in the USGS Hydrological Units #13 (Rio Grande), #14 scale analysis of trends in historical streamflow anomalies (Upper Colorado), and #15 (Lower Colorado) that had in the American Southwest. We modeled past fish species daily streamflow records for 65 or more years (1948– abundance as a function of streamflow anomalies, and 2012) and ≤1% of missing values (n = 120 stations; predicted fish responses 10 years into the future by WebPanel 1). As for the target reaches, we computed combining estimated model coefficients and simulated anomalies during the latest 30-year period of record streamflow anomalies. Methods are described below (see (1983–2012), but using the historical hydrograph (1948– WebPanel 1 for further details). 2012) as a reference. We estimated anomaly trends in target and non-target reaches using non-parametric Time-series datasets Theil-Sen slope estimates and Mann-Kendall trend tests, and bootstrapped 95% confidence intervals (CIs) of We collated time series of streamflow and fish com- Theil-Sen slope distributions to test for the existence of munity composition and abundance for 10 river systems statistically significant basin-wide trends. (16 sites) in the American Southwest. Data originated from several academic, state, and federal agency mon- Fish modeling itoring programs in the Colorado River Basin (Colorado, Gila, San Francisco, San Juan, San Pedro, Verde, and For each river reach, we selected the most common Virgin rivers, and West Clear Creek), Mimbres River subset of species (those occurring in ≥50% of the time Basin, and Rio Grande Basin (Pecos River) (Gido steps) as variates for the Multivariate Autoregressive et al. 2013; JDO unpublished data; Table 1). Fish State-Space (MARSS) model. The three strings of anom- species abundance had been monitored annually for alies were used as covariates at lag 1, representing hy- >10 years, using a standardized protocol within each drological conditions during the 12 months that preceded river system; thus, variation in species abundance in each fish collection. MARSS models are state-space each of the 16 river sites reflected interannual fluc- versions of multivariate autoregressive models that allow tuations in population size. Fish collections occurred for quantifying the effects of environmental drivers on in close proximity to a US Geological Survey (USGS) community dynamics (Holmes et al. 2014). They consist stream-gaging station with historical daily streamflow of a state process and an observation process (see equa- records (≤10% difference in upstream drainage area tions in WebPanel 1), and in this study they were between location of fish collection and gaging station) fitted to obtain the C matrix (whose elements describe (Table 1; WebPanel 1). the effect of each covariate on each species, or streamflow anomaly effects) and the diagonal value of the Streamflow analyses Q matrix (process error variance, describing stochasticity in the fish trajectories). For each river reach, two dif- We parsed out recurrent from stochastic variation in ferent structures were compared: one allowing streamflow flow using signal-processing techniques to describe the anomaly effects to be species-specific (“species model”), streamflow regime at each site. We extracted the sea- and another constraining these effects to be shared among sonal component of streamflow variation in each river all native and all non-native species (“nativity model”). reach using the Discrete Fast Fourier Transform (DFFT) We compared the relative support between the species on normalized, log10-transformed mean daily streamflow and nativity models in a multi-model inference frame- records (1985–2014) (following Sabo and Post 2008). work using the corrected Akaike’s information criterion We computed daily residuals (daily observed – predicted) (AICc). MARSS models were fitted with the “MARSS” over the whole record and then extracted three strings R package (Holmes et al. 2014). of yearly anomalies: the maximum yearly high-flow Subsequently, we simulated stochastic community tra- anomaly (hereafter, “high-flow anomalies”), the mini- jectories and associated species decline risks according to mum yearly low-flow anomaly (hereafter, “low-flow two contrasting future scenarios: (1) “future with trend” anomalies”), and the yearly sum of all daily residuals (streamflow anomalies will continue, over the next 10 (hereafter, “Net Annual Anomalies” [NAA]). High- years, with the same trend observed over the past 30 years and low-flow anomalies represent the magnitude of the [a declining one in most of the rivers]) and (2) “future most extreme events of each year. By contrast, NAA without trend” (streamflow anomalies will stabilize and indexes years with respect to wetter- or drier-than- fluctuate around a stable mean, reflecting conditions that expected (by DFFT), thus measuring the extent and are not worsening) (see schematics in WebFigures 1 and

www.frontiersinecology.org © The Ecological Society of America A Ruhí et al. Declining streamflow and native fish losses

Table 1. Summary characteristics of the studied river reaches and associated MARSS models 467

Drainage area Sites Spp (nat, AICc AICc River, model USGS # (km2) HCDN (years) Period (gaps) non-nat) species nativity Colorado Paria 9382000 293,213 No 1 (13) 2000–2012 (-) 3 (1, 2) 116.95 109.29 (at Lees Ferry, AZ) 9380000 Colorado Havasu 9404200 386,727 No 1 (13) 2000–2012 (-) 8 (3, 5) 291.57 247.75 (above Diamond Cr, AZ) Gila 9430500 4828 Yes 1 (25) 1988–2012 (1) 7 (5, 2) 517.53 493.27 (near Gila, NM) San Francisco 9444000 4281 No 1 (24) 1988–2011 (2) 5 (5, 0) 326.09 298.04 (near Glenwood, NM) San Juan Shiprock 9368000 33,411 No 1 (13) 2001–2013 (-) 11 (4, 7) 482.47 430.24 (at Shiprock, NM) San Juan Four Corners 9371010 37,814 No 1 (13) 2001–2013 (-) 9 (4, 5) 378.53 337.24 (at Four Corners, CO) San Pedro 9471000 3196 No 2 (24) 1990–2013 (-) 6 (2, 4) 886.24 865.88 (at Charleston, AZ) Verde Paulden 9503700 6493 No 1 (15) 1994 – 2008 (-) 9 (3, 6) 402.65 359.43 (near Paulden, AZ) Verde Clarkdale 9504000 9073 No 1 (19) 1994–2012 (-) 8 (3, 5) 450.92 404.55 (near Clarkdale, AZ) Virgin 9415000 13,183 No 2 (16) 1998–2013 (-) 6 (4, 2) 585.44 567.50 (at Littlefield, AZ) West Clear Creek 9505800 624 Yes 1 (11) 2002 – 2012 (2) 3 (1, 2) 120.60 102.14 (near Camp Verde, AZ) Mimbres 8477110 477 No 2 (23) 1990–2012 (-) 5 (4, 1) 594.86 578.81 (at Mimbres, NM) Pecos Acme 8386000 29,474 No 1 (22) 1992–2013 (1) 9 (7, 2) 556.09 531.37 (near Acme, NM) Pecos Lake Arthur 8395500 38,228 No 1 (22) 1992–2013 (-) 9 (6, 3) 594.27 560.73 (near Lake Arthur, NM) Pecos Artesia 8396500 39,627 No 1 (21) 1993–2013 (-) 7 (5, 2) 374.21 364.98 (near Artesia, NM) Pecos Lakewood 8399500 49,632 No 1 (21) 1993–2013 (1) 7 (5, 2) 437.63 406.71 (near Lakewood, NM) Notes: USGS # = code of the USGS stream-gaging station selected to analyze mean daily streamflow with Fourier analyses; Drainage area = drainage area (in km2) at the selected stream-gaging station; HCDN = site being included or excluded from USGS’s 2009 Hydro-Climatic Data Network classification (Lins 2012); Sites (years) = number of sites included in the model as spatial replicates, and number of years with fish abundance data; Period (gaps) = years considered (and number of missing years within the period); Spp (nat, non-nat) = number of modeled species (divided in native, non-native); AICcspecies = support for a MARSS model estimating species-specific trajectories; AICcnativity = support for a MARSS model summarizing species trajectories by nativity status (native versus non-native).

2). Under each scenario, we quantified mean probability extent hydrology limits native viability (favors invasion) (across 1000 realizations) of each fish population crossing across the studied rivers. Finally, we compared the risk an 80% decline threshold over the next 10 years (hereaf- faced by each species between the two near-term scenar- ter, “quasi-extinction risk”). This threshold is relevant for ios (future with trend versus future without trend hydrol- conservation and management because the International ogy), to quantify the unique contribution of streamflow Union for Conservation of Nature (IUCN) considers a anomaly trends to the quasi-extinction risk of each species as critically endangered (according to the Red species (hereafter, “trend effect”). List of Threatened Species) when it demonstrates a decline of 80% over 10 years or three generations (IUCN JJ Results 2013). Using general linear models, we compared quasi- extinction risks between native and non-native species, Streamflow regime properties – describing the mostextreme and we tested the effects of streamflow anomaly trends on high flow of each year (high-flow anomalies), the most risk: if native (non-native) fish face relatively higher extreme low flow of each year (low-flow anomalies), and (lower) risks in river reaches with stronger hydrologic the mean annual flow (NAA) – indicated systemic and alteration, this relationship would quantify to what intensifying hydrologic change. Approximately one-quarter

468 WebFigure 3), implying our study (a) reaches adequately captured the systemic decline in streamflow present across the American Southwest. Streamflow anomalies were robust predictors of past fish abundances: more than one-third (35.7%) of the trajectories of individual species across the modeled rivers (ie in (b) species models) responded significantly to at least one of these spectral metrics of hydrological variation (WebTable 2). Streamflow variation was an even stronger pre- dictor of ensemble trajectories of native and non-native species (ie in nativity models), with close to one- half (45.2%) of these trajectories (c) responding significantly to at least one of the covariates (WebTable 3). Accordingly, nativity models were always better supported than the corresponding species models (Table 1). This suggests that species sharing nativity status (ie native versus non- native fish) have coher- ent trajectories (ie congruent interannual fluctuations in abundance) over time. Figure 1. Trends in streamflow anomalies over the period 1983–2012 across the Native fish faced significantly American Southwest (n = 120) and in the target reaches (n = 16). Overall decreasing higher prospective quasi-extinction trends were observed for (a) high-flow anomalies, (b) low-flow anomalies, and (c) Net risk as compared with non-native Annual Anomalies (NAA). Trends in streamflow anomalies were quantified using non- fish when current hydrological trends are assumed to continue (F1,71 parametric Mann-Kendall trend tests and Theil-Sen slope estimates, and slope 2 distributions were visualized using a combination of violin plots (showing the probability = 4.321, P = 0.041, η p = 0.057 [par- density of the different slopes) and box plots (showing the median, interquartile range tial eta squared, a measure of effect [IQR], and upper/lower whiskers extending from the hinge to the highest/lowest value size]); no such difference was within 1.5 × IQR). We also bootstrapped confidence intervals (CIs) for each found given no future trend in anomalies (F1,71 = 0.338, P = 0.56, distribution of slopes, where the asterisk indicates 95% CIs not overlapping with zero. 2 Inset numbers indicate the proportion of stations with significantly positive (green values) η p = 0.005) (WebTable 6). A or negative (red values) Mann-Kendall trends. model with NAA trend as a covariate confirmed that anomaly trends influence quasi-extinction risks of (n = 32) of the 120 stream-gaging stations across the native and non-native species in opposite directions American Southwest demonstrated statistically signifi when assuming continued hydrological anomalies (F1,69 2 cant negative trends in high-flow anomalies over the = 6.179, P = 0.015, η p = 0.082) (Figure 2a; WebTable 6): past three decades, whereas only two stations showed a streamflow anomaly trends heighten the quasi-extinction significantly positive trend (Figure 1a). The pattern risk for native fish (+8.5% relative risk) while reducing was similar when analyzing low-flow anomalies, with this risk for non- native fish (–5.9% relative risk). This 43 significantly negative and four significantly positive outcome is equivalent to increasing the absolute risk for trends (Figure 1b). Remarkably, NAA displayed signifi- native over non-native fish by 7.4% during the next 10 cantly negative trends for over one-half of the years (Figure 2b). Consistently, more than two-thirds stations (n = 64) and only one significantly posi- (71%; 15 out of 21) of native species demonstrated negative trend in the region, with bootstrapped 95% CIs for tive associations with streamflow anomaly trends, whereas these slope distributions not overlapping with zero less than one-third (29%; 4 out of 14) of non-native spe- (Figure 1c). Overall, estimated slopes in the study reaches cies displayed the same relationship (Figure 2, c and d). were comparable to the regional range (Figure 1; Furthermore, 15 out of the 21 modeled native species had

www.frontiersinecology.org © The Ecological Society of America A Ruhí et al. Declining streamflow and native fish losses higher than a 50% probability of 469 crossing the quasi-extinction (a) (b) threshold in the next 10 years (Figure 2c), with risks being particularly high for four native cyprinids listed as endangered or threatened under the US Endangered Species Act. Despite intraspecific variation in quasi-extinction risk across rivers, only two native species were at low risk of quasi-extinction in at least one of the modeled rivers (WebTable 7).

JJ Discussion

Freshwater fish extinctions have long been associated with climate (c) (d) change and water withdrawals (Xenopoulos et al. 2005; Poff et al. 2012), which, together with invasion, are key components of global change (Staudt et al. 2013). Given the threat that water scarcity poses to freshwater biodiversity globally and particularly in the American Southwest (Vörösmarty et al. 2010; Sabo et al. 2010; Dettinger et al. 2015), anticipating the impacts of climate and hydrological change on Figure 2. Effects of streamflow anomaly trends on quasi-extinction risk of native and imperiled aquatic species is greatly non-native assemblages and species over the next 10 years. (a) In river reaches where needed (McClure et al. 2013). Our streamflow is declining more steeply, native fish face higher quasi-extinction risks than quantitative approach, based on non-natives, and vice versa (lines indicate best linear fits). (b) Accordingly, the trend time-series modeling of fish and flow effect, or mean contribution of streamflow anomaly trends to quasi-extinction risk, data, provided two compelling in- affects native and non-native species in opposite directions (mean ± standard error; the sights about the central role of dashed zero-line represents a hypothetical no trend effect and inset values represent hydrological change on the viability relative changes in quasi-extinction risk caused by anomaly trends). (c and d) Quasi- of native fish populations: (1) extinction risks and trend effects by fish species, conservation category, and nativity status streamflow anomalies explain native (c = native, d = non-native). Circles and error bars show mean, minimum, and and non-native ensemble trajecto- maximum quasi-extinction risk of each species over the next 10 years in the studied river ries, and (2) if dominant downward reaches (potentially ranging from 0 to 1); columns indicate trend effects (potentially trends in streamflow anomalies con- ranging from –1 to +1). Native fish are grouped by status under the US Endangered tinue over the next 10 years (ie Species Act (ESA), and all species are ordered in increasing mean quasi-extinction risk lower extreme high and low flows, within each category. See species codes in WebTable 1. and lower mean annual flow conditions), collapse and replacement of native fish com- rivers are generally assumed to have a competitive munities are to be expected in dryland rivers across advantage over non-native species. However, here we the region. did not model variation in flow per se, but departures Previous studies had reported that alterations to the from the characteristic hydrograph to which the native natural flow regime may favor non-native fish at assemblage is adapted (Bunn and Arthington 2002; the expense of native fish (Bernardo et al. 2003; Lytle and Poff 2004). Trends in these departures – or Propst et al. 2008; Mims and Olden 2013), and our anomaly trends – may represent selection regime modi- results agree with this observation. We further revealed fications, which have been widely recognized as mediat- that streamflow-driven variation in the native versus ing invasion through the removal of direct predators non-native ensembles can be more easily described than and competitors, enhanced supply rate of non-indigenous variation among individual species (as in Ruhí et al. propagules, or the creation of new microhabitats and 2015). Native faunas inhabiting hydrologically variable niches (Byers 2002). We suggest that the lattermost

470 (a) mechanism may be operating in the studied rivers, with some non-native species that characteristically inhabit slow-moving waters (eg families Centrarchidae and Ictaluridae) likely benefitting from muted high flows and “lentification” (Poff et al. 2007; Sabater 2008; Ruhí et al. 2015). Fish species with moderate maturation age, low fecundity, and high investment in juvenile survi- vorship (or equilibrium strategists) have demonstrated the fastest spread in the Lower Colorado River Basin over the past 150 years (Olden et al. 2006). Although life-history theory suggests that these traits should not be favored under frequent high-flow pulses, these conditions that typified dryland rivers in the past are (b) becoming increasingly less common. This study therefore contributes to the notion that increasing changes to dominant streamflow may be decoupling native life histories from their natural flow regimes, while simultaneously favoring some of the most common non-native fish. Emblematic of water resource challenges across much of the American Southwest, the Colorado River is especially vulnerable to water over-exploitation, and the Rio Grande faces large climate-induced deficits in water supply (Dettinger et al. 2015). Flow regulation by dams and over-allocation of resources (via groundwater pumping and surface-water diversions) affect many (c) rivers in these two large river basins, including most of our study sites. The hydrology of the Colorado, San Juan, and Pecos rivers is affected by main-stem impoundments (the primary driver of hydrologic alteration in the US; Carlisle et al. 2010); the free-flowing San Pedro, Upper Verde, and Mimbres rivers are influenced by withdrawal deficits and concomitant decreasing water tables; and only three of the studied sites – San Francisco, Upper Gila, and West Clear Creek – may present close-to-natural flow regimes. The latter two are included in USGS’s 2009 Hydro-Climatic Data Network

(c) G Knowles, USFWS (HCDN; Lins 2012; Table 1), an updated classification Figure 3. (a) Upper Gila River within the Gila Box Riparian of sites where discharge primarily reflects prevailing National Conservation Area near Safford, Arizona. USGS stations meteorological conditions and not human activities (ie in the Upper Gila – including the HCDN station #9430500 – have diversions, storage, or any activity in the drainage basin shown marked negative trends in streamflow over the past 30 years or in the stream channel that may affect the natural (WebTable 5). New Mexico’s Interstate Stream Commission (ISC) flow regime of the watercourse). Interestingly, decreas- plans to divert upstream flow pose further risks to the riverine and ing NAA trends at these reference sites were also statis- riparian ecosystems of this free-flowing section of the river. (b) Verde tically significant, and within the range of the rest of the River at Perkinsville, Arizona, under winter anomalous low flows. sites (WebFigure 3; WebTable 4). This suggests a strong The USGS Northern Arizona Regional Groundwater-Flow Model influence of climate on region-wide patterns of declin- has shown that groundwater pumping in the Prescott region of the ing streamflow, with current and near-future threats to Upper Verde River has reduced the river’s annual baseflow by 4900 the studied rivers likely arising from both altered acre-feet over the past century at the upper end of the Verde Valley, climate and intense river management (Figure 3). with increasing drops expected in the coming decades as human The interaction between climate change, water water demands increase. (c) The spikedace (Meda fulgida) is an resources overuse, and growing demands represents a key endangered cyprinid, historically abundant throughout much of the challenge to water sustainability in the Colorado and Gila River drainage above Phoenix, Arizona. Our study found Rio Grande basins (Dettinger et al. 2015). Despite this, particularly high trend effects for this species, meaning that ongoing opportunities exist to design and prescribe flow regimes hydrological change is exacerbating the quasi- extinction risk of this downstream of dams that will maximize ecological out- species in the river reaches where it can still be found. comes. Intense interdisciplinary work is being done at

www.frontiersinecology.org © The Ecological Society of America A Ruhí et al. Declining streamflow and native fish losses the intersection of water resource engineering and river – that graciously contributed the long-term fish datasets 471 conservation ecology (Acreman et al. 2014; Poff et al. to JDO that were analyzed in this study. We also 2015), and screening frameworks could help prioritizing thank K Fritschie for assisting in database management, dams where engineered flows may be warranted and fea- and members of the Sabo lab for suggestions that sible (Grantham et al. 2014). Although dams can facili- improved the manuscript. Support for this work was tate the establishment and spread of non-native fish at provided by the US National Science Foundation the expense of locally adapted native faunas (Johnson (1204478) to AR and JLS, and by the US Department et al. 2008; Mims and Olden 2013), operating dams to of Defense SERDP (RC-2511) to JDO. recreate seasonal high-flow pulses or to mitigate extreme low flows in drought years may often be feasible, allow- JJ References ing a balance to be established between socioeconomic Acreman M, Arthington AH, Colloff MJ, et al. 2014. Environmental flows for natural, hybrid, and novel riverine goods and services and ecosystem needs (Poff et al. 2003, ecosystems in a changing world. Front Ecol Environ 12: 466–73. 2015). Large-scale flow experiments have shown promis- Bernardo JM, Ilhéu M, Matono P, and Costa AM. 2003. ing results in many rivers (Shafroth et al. 2010; Olden Interannual variation of fish assemblage structure in a et al. 2014), and our study suggests that restoring Mediterranean river: implications of streamflow on the seasonal high-flow pulses through dam operations in a dominance of native or exotic species. River Res Appl 19: 521–32. sustained manner – every year, wherever possible – could Bunn SE and Arthington AH. 2002. Basic principles and ecologi- contribute to the long-term persistence of many native cal consequences of altered flow regimes for aquatic biodiver- fish species. In free-flowing watercourses, increasing the sity. Environ Manage 30: 492–507. frequency and magnitude of high-flow events may not be Byers JE. 2002. Impact of non-indigenous species on natives possible, but protecting base flows by pausing diversion enhanced by anthropogenic alteration of selection regimes. Oikos 97: 449–58. or groundwater pumping operations during supraseasonal Carlisle DM, Wolock DM, and Meador MR. 2010. Alteration of droughts could offset climate-driven low-flow anomaly streamflow magnitudes and potential ecological consequences: trends. This adaptive management strategy would be a multiregional assessment. Front Ecol Environ 9: 264–70. especially beneficial when hydrological connectivity is Dettinger M, Udall B, and Georgakakos A. 2015. Western water at risk, and aseasonal low flows threaten the dispersal of and climate change. Ecol Appl 25: 2069–93. Gido KB, Propst DL, Olden JD, et al. 2013. Multidecadal responses of fish across the river networks (Jaeger et al. 2014). Given native and introduced fishes to natural and altered flow regimes in that return to reference flows is no longer feasible in the American Southwest. Can J Fish Aquat Sci 70: 554–64. many rivers in the American Southwest, a “designer” Grantham TE, Viers JH, and Moyle PB. 2014. Systematic screen- approach (Acreman et al. 2014) focused on flow anoma- ing of dams for environmental flow assessment and implemen- lies may be a feasible way to manage for native species tation. BioScience 64: 1006–18. Holmes ER, Ward E, and Wills K. 2014. Package “MARSS”. R persistence. package version 3.9. http://CRAN.R-project.org/package =MARSS. Viewed 22 Aug 2016. JJ Conclusions IPCC (Intergovernmental Panel on Climate Change). 2014. Climate change 2014: impacts, adaptation, and vulnerability. Cambridge, UK, and New York, NY: Cambridge University Press. Increasing evidence suggests that climate and hydro- IUCN (International Union for Conservation of Nature). 2013. logical change need to be considered when assessing Guidelines for using the IUCN Red List categories and criteria, risk to aquatic endangered species (McClure et al. Version 10. Gland, Switzerland: IUCN. 2013). We show that variation in streamflow controls Jaeger KL, Olden JD, and Pelland NA. 2014. Climate change poised to threaten hydrologic connectivity and endemic fishes native and non-native fish abundances in the American in dryland streams. P Natl Acad Sci USA 111: 13894–99. Southwest, and the near-term fate of several species Johnson PTJ, Olden JD, and Vander Zanden MJ. 2008. Dam of conservation interest. Predicted quasi-extinction risks invaders: impoundments facilitate biological invasions into are high (>50%) for nearly three-quarters of the mod- freshwaters. Front Ecol Environ 6: 357–63. eled native species, and current trends in streamflow Lake PS. 2003. Ecological effects of perturbation by drought in flowing waters. Freshwater Biol 48: 1161–72. anomalies increase risk for natives but reduce risk for Lake PS. 2011. Drought and aquatic ecosystems: effects and non-natives. Species invasion and replacement of native responses. Chichester, UK: Wiley-Blackwell. fish is therefore an emergent outcome of widespread Lins HF. 2012. USGS Hydro-Climatic Data Network 2009 flow alteration, and continued homogenization of fish (HCDN-2009). Reston, VA: US Geological Survey. communities is very likely in the near future unless Lytle DA and Poff NL. 2004. Adaptation to natural flow regimes. Trends Ecol Evol 19: 94–100. streamflow anomaly trends can be minimized. McClure MM, Alexander M, Borggaard D, et al. 2013. Incorporating climate science in applications of the US Endangered Species JJ Acknowledgements Act for aquatic species. Conserv Biol 27: 1222–33. Mims MC and Olden JD. 2013. Fish assemblages respond to altered We thank the numerous individuals, agencies, and flow regimes via ecological filtering of life history strategies. Freshwater Biol 58: 50–62. institutions – including the Arizona Game and Fish Olden JD, Konrad CP, Melis TS, et al. 2014. Are large-scale flow Department, the New Mexico Department of Game experiments informing the science and management of fresh- and Fish, the Navajo Nation, and the US Forest Service water ecosystems? Front Ecol Environ 12: 176–85.

472 Olden JD, Poff NL, and Bestgen KR. 2006. Life-history strategies Sabo JL. 2014. Predicting the river’s blue line for fish conservation. predict fish invasions and extirpations in the Colorado River P Natl Acad Sci USA 111: 13686–87. Basin. Ecol Monogr 76: 25–40. Sabo JL and Post DM. 2008. Quantifying periodic, stochastic, Perkin JS, Gido KB, Costigan KH, et al. 2015. Fragmentation and and catastrophic environmental variation. Ecol Monogr 78: drying ratchet down Great Plains stream fish diversity. Aquat 19–40. Conserv 25: 639–55. Sabo JL, Sinha T, Bowling LC, et al. 2010. Reclaiming freshwater Poff NL, Allan JD, Palmer MA, et al. 2003. River flows and water sustainability in the Cadillac Desert. P Natl Acad Sci USA 107: wars: emerging science for environmental decision making. 21263–70. Front Ecol Environ 1: 298–306. Seager R, Ting M, Li C, et al. 2013. Projections of declining Poff NL, Brown CM, Grantham TE, et al. 2015. Sustainable surface-water availability for the southwestern United States. water management under future uncertainty with Nat Clim Change 3: 482–86. eco- engineering decision scaling. Nat Clim Change; Shafroth PB, Wilcox AC, Lytle DA, et al. 2010. Ecosystem effects doi:10.1038/nclimate2765. of environmental flows: modelling and experimental floods in a Poff NL, Olden JD, Merritt DM, and Pepin DM. 2007. dryland river. Freshwater Biol 55: 68–85. Homogenization of regional river dynamics by dams and Staudt A, Leidner AK, Howard J, et al. 2013. The added compli- global biodiversity implications. P Natl Acad Sci USA 104: cations of climate change: understanding and managing 5732–37. biodiversity and ecosystems. Front Ecol Environ 11: 494–501. Poff NL, Olden JD, and Strayer DL. 2012. Climate change and Vörösmarty CJ, McIntyre PB, Gessner MO, et al. 2010. Global freshwater fauna extinction risk. In: Hannah L (Ed). Saving a threats to human water security and river biodiversity. Nature million species: extinction risk from climate change. 467: 555–61. Washington, DC: Island Press. Xenopoulos MA, Lodge DM, Alcamo J, et al. 2005. Scenarios of Propst DL, Gido KB, and Stefferud JA. 2008. Natural flow regimes, freshwater fish extinctions from climate change and water nonnative fishes, and native fish persistence in arid-land river withdrawal. Glob Change Biol 11: 1557–64. systems. Ecol Appl 18: 1236–52. Ruhí A, Holmes EE, Rinne JN, and Sabo JL. 2015. Anomalous JJ droughts, not invasion, decrease persistence of native fishes in Supporting Information a desert river. Glob Change Biol 21: 1482–96. Sabater S. 2008. Alterations of the global water cycle and their Additional, web-only material may be found in the effects on river structure, function and services. Freshwater Rev online version of this article at http://onlinelibrary. 1: 75–88. wiley.com/doi/10.1002/fee.1424/suppinfo