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Ecological Indicators 45 (2014) 456–464

Contents lists available at ScienceDirect

Ecological Indicators

jo ur nal ho me page: www.elsev ier .com/ locate / ecol ind

esponses of Mediterranean aquatic and riparian communities touman pressures at different spatial scales

aniel Brunoa,∗, Oscar Belmarb, David Sánchez-Fernándezc, Simone Guareschia,ndrés Millána, Josefa Velascoa

Departamento de Ecología e Hidrología, Facultad de Biología, Universidad de Murcia, Campus de Excelencia Internacional Regional ‘Campus Mareostrum’, 30100 Murcia, SpainIrstea, UR HYAX Hydrobiologie, HYNES (Irstea – EDF R&D), F-13182 Aix-en-Provence, FranceInstitut de Biologia Evolutiva (IBE, CSIC-UPF), Passeig maritim de la Barceloneta 37-49, 08003 Barcelona, Spain

r t i c l e i n f o

rticle history:eceived 18 February 2014eceived in revised form 29 April 2014ccepted 30 April 2014

eywords:reshwater ecosystemsnthropogenic disturbancecale-dependenceacroinvertebrates

iparian vegetation

a b s t r a c t

Mediterranean river ecosystems are subjected to intense human pressures and impacts that affect boththeir aquatic and riparian communities. However, given their stratified position in the river ecosys-tem and varying ecological requirements, aquatic and riparian communities can respond differently tosuch pressures. These biological responses could also vary depending on the nature of the disturbances,the spatial scale considered and the indicators used as response variable. Here, we aim to assess theinfluence of the main human pressures present in Mediterranean rivers (agricultural land use and hydro-morphological alteration) on the biodiversity and ecological condition indicators of both riparian andaquatic communities at two spatial scales: reach and basin. For this purpose, a total of 56 samplingsites covering the study area (Segura basin, SE Spain) were surveyed. Water beetles and woody riparianvegetation richness were used respectively as biodiversity surrogates of aquatic and riparian commu-nities, and the Iberian Biomonitoring Working Party (IBMWP) and Riparian Quality Index (RQI) wereused to assess the ecological status of both communities. As expected, we found a general decrease inboth richness and ecological condition when human pressures increased, regardless of the spatial scaleconsidered. Nonetheless, agricultural land use was the main pressure explaining riparian richness andquality, whereas aquatic communities’ responses were highly related to hydromorphological alteration.

Contrary to expected, in general, variables at basin scale had a greater effect than those variables oper-ating at local scale. In addition, ecological condition indices responded more clearly to human pressuresthan biodiversity surrogates. Therefore, land use and hydrological planning at basin scale are essentialcomplements to conservation and restoration efforts, traditionally carried out at reach scale, in order tomaintain stream ecosystem integrity and biodiversity.

© 2014 Elsevier Ltd. All rights reserved.

. Introduction

The high human pressures on inland aquatic ecosystems haveaused these habitats to become recognised as some of the mosthreatened in the world (Saunders et al., 2002). This is especially

vident in the Mediterranean Basin, one of the Earth’s biodi-ersity hotspots (Myers et al., 2000), where the long history ofubstantial human impacts on the landscape and fluvial systems

∗ Corresponding author. Tel.: +34 868884977; fax: +34 868883963.E-mail addresses: dbrunocollados@um.es (D. Bruno),

scar.belmar-diaz@irstea.fr (O. Belmar), davidsan@um.esD. Sánchez-Fernández), simone.guareschi@um.es (S. Guareschi), acmillan@um.esA. Millán), jvelasco@um.es (J. Velasco).

ttp://dx.doi.org/10.1016/j.ecolind.2014.04.051470-160X/© 2014 Elsevier Ltd. All rights reserved.

has been well-documented (Hooke, 2006). These areas provide awide anthropogenic gradient suitable for studying the influenceof human pressures on fluvial communities. Thus, understandingthe processes and relationships between human pressures, spatialheterogeneity and riverine communities in Mediterranean areasconstitutes a major challenge in freshwater ecology and conserva-tion (Bonada and Resh, 2013; Cooper et al., 2013), particularly inthe current context of global change (Sala et al., 2000).

Although frequently overlooked in limnological studies(Ferreira and Aguiar, 2006), riparian areas are an integral partof riverine ecosystems (Ward et al., 2002) that influence both

the structure and functioning of aquatic communities (Sabateret al., 2000). Thus, aquatic and riparian communities are mutuallyinterdependent in terms of ecological processes. These largeterrestrial-aquatic linkages occur in both directions, existing

l Indicators 45 (2014) 456–464 457

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mportant flows of energy and material between these adjacentystems (Naiman and Décamps, 1997). Conservation and man-gement strategies are usually focussed on aquatic or riparianommunity but sparsely on both of them. Despite the existingnterconnection between both communities, given their stratifiedosition in the river ecosystem and their varying functional traitsnd ecological requirements, their responses to human pressuresould also differ.

Riparian and aquatic communities’ responses could also differepending upon the nature of these pressures and the spatial scaleonsidered (Aguiar et al., 2009; Ferreira and Aguiar, 2006; Richardst al., 1996). The hydromorphological alteration of rivers as aesult of flow regulation, and land-use changes can be consideredmong the main human disturbances modifying freshwater biolog-cal communities, especially in Mediterranean areas (Belmar et al.,013; Bruno et al., 2014; Stella et al., 2013), where they exacerbatehe natural stress of river ecosystems (Kroll et al., 2013; Strombergt al., 2004). Given the scarcity of water resources and the widegricultural surfaces in these areas, both pressures are highly inter-onnected, as water supply for irrigation is the main cause forater demands (Zimmer, 2010). Therefore, although agriculture

nd hydromorphological alteration cause deep modifications inquatic and riparian communities (Allan, 2004; Ward, 1998), theiresponses to these pressures can differ depending on the nature ofhe human pressure. We consider that agricultural land use coulde exerting a greater influence on riparian communities, since theyre directly impacted by agricultural processes (e.g. occupation ofiparian area, greater amount of sediments, modification of riverank profile), whereas strictly aquatic communities could respondore clearly to in-stream disturbances such as hydromorpholo-

ical alteration (e.g. low ability to persist after the hydropeakingrocess caused by dam releases, micro and mesohabitat reductionnd homogenisation, or bed substrate modification).

Regarding the spatial scale, most previous studies have reportedhat land-use and hydrological alteration acting at local scalere important for shaping river communities (e.g. Fernandest al., 2011; Nerbonne and Vondracek, 2001; Nilsson et al., 1991;ponseller et al., 2001), and some of them have also documentedhe importance of these pressures at wider scales (Wahl et al., 2013;ates and Bailey, 2011). Thus, although alteration processes at bothcales seem to be important (Johnson et al., 2007; Marzin et al.,013; Stewart et al., 2001), a greater influence of those operat-

ng at reach scale can be detected, as both aquatic and riparianommunities are directly exposed to these local impacts, espe-ially in semi-arid rivers (Aguiar and Ferreira, 2005; Boyero, 2003;onteagudo et al., 2012).Finally, as results can also vary depending on the indicator used

o estimate the biological communities’ responses, we decided tose both biodiversity and ecological condition indicators. Firstly,iodiversity indicators are well-surveyed and taxonomically stableroups of organisms whose species richness patterns or rarity cane considered as similar to those of unsurveyed taxa in the sameegion (Pearson, 1994). Secondly, indicators of ecological conditionre widely used to assess river health (Bonada et al., 2006; Karr,999). They are complementary integrative and holistic indices that

nvolve a wide range of metrics in order to obtain a measure ofuality (Feest et al., 2010). Although it seems evident that ecolog-

cal condition indices should clearly respond to human pressures,s they are ultimately designed to assess their effects, this assump-ion has been scarcely tested. Furthermore, many ecological studiesegarding the influence of human pressures on biological commu-ities continue using richness as a response variable (Birk et al.,

012).

Many studies have separately tested the influence of diverseuman pressures on river communities, the varying biologi-al responses depending on the spatial scale considered, the

Fig. 1. Geographic location of the study area showing the 56 sampling sites, agri-cultural area and the main dams.

convergence on the sensitivity of several ecological indicators orthe contrasting response of aquatic and riparian communities fac-ing these anthropogenic disturbances. Nevertheless, as far as weknow, no holistic studies considering all of these different aspectstogether (kind of pressure, scale, response variable and target com-munity) have been conducted. Therefore, in this study, we aim toassess the influence of agricultural land use and hydromorpholo-gical alteration at basin and reach scales on the biodiversity andecological condition of both Mediterranean riparian and aquaticcommunities. Based on the information above, we predict the fol-lowing: (1) although both types of human pressures are expectedto modify riparian and aquatic communities, agricultural land usecould have a greater influence on riparian communities, whereashydromorphological alteration could have a more clear relation-ship with strictly aquatic communities; (2) human pressures actingat reach and basin scales will modify riparian and aquatic commu-nities, although a greater influence from those operating at reachscale may be found; (3) ecological condition indicators will be moresensitive to human pressures in both communities than biodiver-sity indicators.

2. Methods

2.1. Study area

Located in one of the most arid zones in Europe (Fig. 1), theSegura River Basin is an environmentally diverse basin due tohuman (alteration) and natural (climatic) gradients with a predo-minately semi-arid climate. These marked gradients make theman ideal candidate to be utilised as a Mediterranean pilot basin, asthey involve a wide range of conditions that can be present in most

Mediterranean basins. In general, the river network ranges fromlowly populated forested headwaters to densely populated low-land cities, with a mostly shrubby landscape. Despite the notablepresence of forested or semi-natural areas (45.2%), particularly in

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he headwaters, the intense expansion of land dedicated to agricul-ural use (52.1%) (estimated from Corine Land Cover 2000) duringecent decades has caused a progressive reduction of natural andemi-natural areas. In addition, water demand exceeds 224% of thatvailable and only 4% of runoff reaches the mouth of the Seguraiver (Zimmer, 2010). As one of the most regulated river networks

n Europe (CHS, 2007; MMA, 2004), a high density of dams (24 largeams, an average of one every 50 km of channel) counteracts thishortage in water resources. Accordingly, agricultural pressuresnd dam regulation have been recognised as two of the most impor-ant spoilers of aquatic and riparian ecosystems in the study areaBelmar et al., 2010; Bruno et al., 2014; Kroll et al., 2013).

.2. Data collection

.2.1. Biological dataA total of 56 sampling sites were selected to represent the

ide range of human pressures present in the basin (Fig. 1), thatimultaneously minimise the confounding effect of the main nat-ral stressors (i.e. water salinity and flow temporality). For thiseason, we selected permanent flow rivers with electrical conduc-

ivity < 5000 �S cm−1. Finally, urban areas were excluded becausehey represent very particular conditions that present an isolatedistribution and cover only 2.1% of the entire study area (estimatedrom Corine Land Cover 2000).

ig. 2. Distribution of riparian and macroinvertebrate quality and richness values in the stpecies richness.

ators 45 (2014) 456–464

Woody riparian vegetation richness and the Riparian QualityIndex (RQI, González del Tánago et al., 2006) were used as sur-rogates of biodiversity and to assess the ecological status of theriparian communities, respectively. RQI is a widely used index inthe Iberian Peninsula because it has been proven to detect humanimpacts on riparian ecosystems (Barquín et al., 2011; Belmar et al.,2013; Navarro-Llácer et al., 2010). RQI is composed of several sub-indices characterising lateral, longitudinal and vertical continuity,composition, structure, regeneration and bank condition. Althoughthe effect of anthropogenic alteration on vegetation could varydepending upon the function and features of the type of vegetationconsidered (Bunn and Arthington, 2002), woody riparian speciesare long-lived, stable and sensitive to human pressures, and cantherefore be used as suitable indicators of hydromorphological andland-use disturbances (Bejarano et al., 2012; Nilsson et al., 1991;Villarreal et al., 2012).

Coleoptera richness and the Iberian Biomonitoring WorkingParty (IBMWP, Alba-Tercedor et al., 2002) were used as a surro-gate of biodiversity and an indicator of the ecological status forthe aquatic communities, respectively. Water beetles richness wasused as a surrogate of biodiversity for strictly aquatic communi-ties, as it has proven showed to be a good biodiversity indicator in

aquatic ecosystems (Guareschi et al., 2012), especially in the studyarea (Sánchez-Fernández et al., 2006). Their high diversity makesthem potential candidates for assessing the effects of land use, evenat larger spatial scales (Minaya et al., 2013). The IBMWP is a quality

udy area: (a) RQI, (b) Riparian woody species richness, (c) IBMWP and (d) Coleoptera

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ndex largely used in the Iberian Peninsula (Munné and Prat, 2009),ased on differential tolerance to the organic pollution of aquaticacroinvertebrate families. Although both biological indicators

an be positively correlated, the use of a higher taxonomic reso-ution (water beetle species richness) could provide additional andomplementary information to family based indices such as IBMWPn terms of ecological response to disturbances (Kroll et al., 2013).

Riparian data were obtained for each sampling site along 500 mength reaches, whereas macroinvertebrates were only collectedn the last 100 m (downstream). Riparian and macroinvertebrateurveys were conducted between 2010 and 2011 during late springnd early summer, corresponding to the maximum species activityeason (invertebrates) and also the most suitable period for riparianegetation, particularly when using one-shot surveys (Ferreira andguiar, 2006). Furthermore, it has been observed that this seasonalample taken from Mediterranean streams is representative of theooled annual macroinvertebrate community (Bonada et al., 2007).

Within each survey unit, we recorded the occurrence of woodyiparian species, through ten transects from the water margin up tohe natural bankfull width, obtaining a unique list of species by site.dditionally, the RQI index was also calculated for each samplingite. Macroinvertebrates were sampled with a kick-net of 500 �mesh, following a multihabitat standardised protocol, covering allesohabitat types present in the reach (Jáimez-Cuéllar et al., 2002).

ach kick-sample was examined in the field and sampling sitesere consistently surveyed until no further families or water beetle

pecies (morphotypes) were found in each mesohabitat. The kick-ample contents were pooled into a unique site-sample, preservedn 96% ethanol and identified to family level, except for Coleoptera,

hich were identified in the laboratory to species level in order tobtain the water beetle richness and IBMWP score for each locality.

.2.2. Human pressures at reach and basin scalesAt basin scale, a hydromorphological alteration score was

erived using the number of dams (count), their regulatory capac-ty (hm3) and the area of irrigated land (%) (as a surrogate of waterxtraction) draining to each sampling site, as these variables aressociated with the main hydromorphological alterations in thetudy area (see Belmar et al., 2013). At reach scale, hydromor-hological alteration was assessed using the habitat modificationcore (HMS) derived from the River Habitat Survey protocolEnvironment Agency, 2003; Raven et al., 1997), which assesses theydromorphological modification of the river channel in the 500 meach, considering weirs, dams, outfalls, realignments, reinforce-ent and resectioning among other variables. It has been widely

sed as a surrogate of hydromorphological alteration (e.g. Bonat al., 2008; Erba et al., 2006).

The percentage of agriculture (both irrigated and non-irrigatedrops) at reach (200 m buffer at both sides of the sampling site; i.e.n area of 200,000 m2) and basin (entire area draining to the samp-ing site) scales were computed for each locality by combining thercGIS software 9.2 (ESRI, Redlands, California, USA) and the analy-is toolkit NetMap (Benda et al., 2007), based on the available digitalnformation (1:25,000) provided by the Occupation Informationystem of Soil in Spain (SIOSE).

.3. Statistical analyses

Spearman’s rank correlation coefficients were used to discardighly (R > 0.7) correlated human pressures at the different spa-ial scales considered to reduce collinearity among predictors.lthough the anthropogenic variables were moderately corre-

ated, none of them showed a Spearman correlation coefficientigher than 0.7, and all of them were used in the analyses. Gen-ralised Linear Models (GLMs) (McCullagh and Nelder, 1989) withifferent combinations of variables (a total of 36 models) were

ators 45 (2014) 456–464 459

performed to test the influence of the (1) individual anthropogenicvariables [a: Reach agriculture (200 m buffer), b: Basin agriculture,c: Reach hydromorphological alteration (HMS), d: Basin hydro-morphological alteration], (2) type of pressure (agriculture: a + b;hydromorphological alteration: c + d), (3) spatial scale (reach: a + c;basin: b + d) and (4) the combination of all types of pressures andscales considered (a + b + c + d) (see Appendix A, Fig. S1) on riparianand aquatic ecological condition (RQI and IBMWP) and species rich-ness (riparian and water beetle richness). GLMs were carried outassuming a Poisson distribution for the dependent variables. A step-wise procedure was used to insert environmental variables into themodel (Nicholls, 1989) and linear and quadratic relationships weretested. The percentage of explained deviance was used to evaluatethe models’ performance. Outliers were identified and removed ineach model (if applicable). Finally, normality (Shapiro–Wilk test)and spatial independence (Moran’s I test, ArcGIS 9.2) of residualswere assessed. Prior to the analyses, hydromorphological variables(c and d) were log-transformed and agricultural percentages at bothscales (a and b) were arcsine-square-root transformed to improvelinearity. All explanatory variables were z-standardised (mean = 0,SD = 1) to allow model coefficient comparison. These statisticalanalyses were carried out using the ‘corrgram’, ‘car’ and ‘MASS’packages of the statistical computing software R (R DevelopmentCore Team, 2012).

3. Results

Strong decreasing patterns from rich, well-conserved headwa-ters to poor, impaired lowlands were found for both riparian andmacroinvertebrate communities (Fig. 2). The values of agriculturalarea ranged from 0 to 67.8% at reach scale and from 0 to 94.6%at basin scale. The hydromorphological alteration at reach scale(HMS) ranged from 0 to 3480, whereas the basin hydromorpholo-gical index at basin scale ranged from 0 (minimum flow alteration)to 14 (maximum flow alteration) among the different localities.

In general, GLMs showed a clear decrease in richness andecological condition for both riparian (Fig. 3) and aquatic macroin-vertebrate communities (Fig. 4) when human pressure increased,independently of the type of pressure or the spatial scale consid-ered (Table 1). It is worth noting that, with some exceptions (e.g.linear response to basin agriculture) most of the relationships werehumped. This general decrease takes place with the exception of thefirst phases of alteration, in which species richness and ecologicalcondition seem to remain constant or even, in some cases, increaseslightly.

Although the patterns were similar for riparian and aquaticmacroinvertebrates, differences in the percentage of explaineddeviance were detected (Table 1). For example, in general, thegreater percentages of explained deviance corresponded to agricul-tural land use for riparian communities and hydromorphologicaldisturbances for macroinvertebrates. The basin scale pressureswere able to explain a higher percentage of deviance than thoseat local scale. Moreover, it seems that ecological condition indi-cators responded more clearly than biodiversity indices to humanpressures. Lastly, models combining all of the pressures displayeda relatively high percentage of explained deviance, especially forriparian and macroinvertebrate-based condition indices.

All of the models’ residuals showed spatial independence,whereas some exceptions were found in the normal distributionassumption (6 of 36 models showed p-values between 0.01 and

0.05). However, the risk of false significant relationships in GLMresults is minimal given the high number of samples and the highsignificance of the variables and terms entered into the models(p < 0.01) (see Appendix A, Table S1).

460 D. Bruno et al. / Ecological Indicators 45 (2014) 456–464

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ig. 3. Response of riparian condition (RQI) and riparian richness to agriculturend hydromorphological disturbances at reach and basin scales according to GLMesults.

. Discussion

We found a general decrease in species richness and the eco-ogical quality of both riparian and aquatic communities from theeadwaters to the lowlands. This clear negative relationship was

able 1LM analyses showing the percentages of explained deviance for significant (p < 0.01)

ressures.

Anthropogenic disturbance RQI Rip

Reach agriculture (a) 51.4 40.Basin agriculture (b) 34.3 45.Reach hydromorphological alteration (c) 39.8 22.Basin hydromorphological alteration (d) 41.4 29.Agriculture (a + b) 54.6 51.Hydromorphological alteration (c + d) 48.0 29.Reach scale (a + c) 63.7 45.Basin scale (b + d) 50.1 50.Whole (a + b + c + d) 67.4 55.

Fig. 4. Response of macroinvertebrate-based quality index (IBMWP) and water bee-tle richness to agriculture and hydromorphological disturbances at reach and basinscales according to GLM results.

found for both agricultural land use and hydromorphological alter-ation regardless of the spatial scale considered, thereby confirming

the importance of human pressures shaping biological communi-ties in fluvial ecosystems, and particularly in Mediterranean rivers.The similarity in the responses displayed by both communitiesagrees with results obtained in other climatic areas (Rios and Bailey,

individual variables, type of impact, scale and the combination of all considered

arian richness IBMWP Coleoptera richness

0 41.1 26.16 51.4 45.46 60.9 39.21 64.5 54.21 60.2 45.41 70.0 54.20 74.6 44.86 78.5 53.40 82.9 57.9

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006), evidencing that both communities are intensively inter-onnected through physical processes and fluxes of energy andutrients, and respond similarly to multi-scaled abiotic driversGregory et al., 1991).

Although a general decline in richness and ecological conditionas observed in both communities as human pressures intensi-ed, at low rates of alteration the values of both types of biological

ndicators remained constant or even increased slightly, partic-larly regarding hydromorphological alteration. Opposite to theetrimental and impoverishing effect of medium or high intensityisturbances on target communities, low-intensity stream mod-

fication could promote certain habitat heterogeneity (Bertrandt al., 2004), leading to a slight increase in macroinvertebrate andiparian-based values. Another possible explanation is the highariability (data dispersion) of index values found in survey unitst low alteration rates. As several human activities can be simul-aneously altering riverine communities, this dispersion could beue to the effects of other human pressures or natural stressors.aking into account that we are only using anthropogenic vari-bles, models combining all the pressures displayed a relativelyigh percentage of explained deviance.

.1. Type of impact: agriculture vs. hydromorphological alteration

Although riparian and aquatic communities were negativelynfluenced by both types of human pressures, in general, agri-ultural land use seemed to be the most important disturbanceor riparian communities, whereas hydromorphological alterationhowed a clearer relationship to aquatic communities. Therefore,espite the dramatic ecological implications that stream modifi-ation has on the structure and function of river ecosystems as

whole (Elosegi and Sabater, 2013), agriculture seems to exert aore intense effect on riparian vegetation, since it is more directly

nd physically impacted by some of these agricultural practices.griculture constitutes an impact that not only constrains the ripar-

an area by causing the loss of riparian forests, it also alters river andiparian integrity, habitat quality, bank stability nutrient enrich-ent and the sedimentation process (Allan, 2004; Roth et al., 1996).

inally, both types of disturbance favour the invasion of weedsnd exotic, opportunistic species and a decline in woody speciesFerreira and Moreira, 1995; Greet et al., 2012). This species sub-titution usually involves a modification of the functional traitombination of the riparian community, which can lead to a reduc-ion of the ecosystem services it provides (e.g. soil fixation andrganic matter supply).

The more direct relationship between hydromorphologicallteration and strictly aquatic communities reflects the lowerbility of freshwater biota to withstand this disturbance thatmpoverishes fluvial communities. Previous analyses in the studyrea have shown the important role of natural flow variability inetermining the composition and richness of macroinvertebratesBelmar et al., 2012). Although aquatic Mediterranean communi-ies have evolutionary mechanisms and adaptations to face thextreme natural events and predictable flow variability typical ofhese semi-arid areas (e.g. flash floods and droughts), they can-ot cope with those caused by the anthropogenic alteration ofow regimes, given their different timing, frequency, predictabil-

ty and magnitude (Kroll et al., 2013). Its negative effects can beven worse if such altered flow discharge occurs during sensitiveife stages, as is the case with large-scale, unpredictable and fre-uent dam releases during summer to meet agricultural demands.nly resistant and rapidly dispersive species are able to avoid their

ffects (Allan, 2004). Rivers and streams in the study area havexperienced significant changes in flow regimes by dam regulationimilar to those described in other arid and semiarid Mediter-anean areas: the dams have reduced the flow magnitude, changed

ators 45 (2014) 456–464 461

their frequency and inverted seasonal flow patterns downstream(Grantham et al., 2013). These changes have resulted in increasedchannel dimensions, homogeneous aquatic habitats, altered sed-imentation and channel-bed grain size, as well as the absence ofin-channel debris and submerged vegetation (Bunn and Arthington,2002; Ligon et al., 1995). This reduces the diversity of resourcesand refuges for fauna and has negative effects on banks and ripar-ian vegetation too (Belmar et al., 2013; Navarro-Llácer et al., 2010).Hydromorphological alteration can also produce an increase in theduration of droughts in the most arid areas, leading to the “ter-restrialisation” of fluvial ecosystems and loss of freshwater biota(Sabater and Tockner, 2010).

The weaker relationship between aquatic community and agri-culture in comparison with hydromorphological alteration couldbe due to the buffering effect of riparian areas mitigating the effectof agricultural land-use in the basin, avoiding a deeper impact onaquatic communities, particularly in areas hosting well-conservedriparian galleries (Naiman and Décamps, 1997; Riseng et al., 2011).Thus, further research will be necessary to unravel the potential roleof riparian vegetation in mitigating the impact produced by humanpressures on aquatic communities (i.e. not as a variable response,but as a predictor).

4.2. Spatial Scale dependence

Our results showed that human pressures negatively influencedaquatic and riparian communities, regardless of the scale consid-ered, as pointed out by previous studies (Allan, 2004; Burcheret al., 2007; Gregory et al., 1991). We expected a clearer relation-ship between riverine communities and human pressures actingat reach scale than those at basin scale, as other authors hadpreviously found (Moerke and Lamberti, 2006: Sponseller et al.,2001; Strayer et al., 2003). In fact, several studies did not find rela-tionships between catchment disturbances (land cover) and rivercommunities (Heino et al., 2002; Rios and Bailey, 2006). However,our results did not meet our predictions. We found that variablesat basin scale had an effect at least comparable or even greaterto those operating at the local scale. These findings agree withother studies stating that human pressures acting at basin scaleseem to also play a major role in riparian vegetation (Aguiar et al.,2009; Salinas and Casas, 2007) and macroinvertebrate communi-ties (Dolédec et al., 2011; Magierowski et al., 2012). Therefore, anon-locally impacted reach does not necessarily result in good eco-logical condition or biodiversity values, as catchment disturbancescould exert a strong influence through erosion, sediment load,chemistry (Törnblom et al., 2011), organic matter input, in-streamprimary production (Dolédec et al., 2011), the invasion of exoticspecies (Greet et al., 2012; Poff et al., 2007) or habitat homogeni-sation (Belmar et al., 2013). Furthermore, large-scale and chronicdisturbances in a watershed can limit river diversity over the longterm (Harding et al., 1998). However, the majority of restorationefforts are focussed on reducing local alteration, which is insuffi-cient in order to guarantee acceptable ecological condition valuesin freshwater ecosystems (Bernhardt and Palmer, 2011; Kail andWolter, 2013). Thus, this restoration effort should focus on meas-ures at the river basin scale in a greater extent in order to developmore effective management strategies.

4.3. Type of indicator: biodiversity vs. ecological condition

Although we found a high congruence in the indicators’response to human disturbances, as expected, the sensitiveness of

ecological condition indicators (RQI and IBMWP) were generallyhigher than that of those for biodiversity (richness of woody ripar-ian and water beetles), given their holistic and integrative nature.Despite the methodological differences between the ecological

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ondition indicators used here, they have been identified as sen-itive to different types of disturbances including land use changend stream modification (Belmar et al., 2013; Garófano-Gómezt al., 2013; Munné and Prat, 2009; Sánchez-Montoya et al., 2010).n particular, among all the indicators used here, IBMWP washe most sensitive one. It seems that its scoring system based onhe different tolerance to pollution of aquatic macroinvertebrateamilies could be sensitive even for no pollutant human pressuresi.e. hydromorphological alteration).

However, the lower sensitivity of richness measures can be dueo the fact that human pressures produced an impoverishment orimplification in the communities, modifying not only the numberf species, but also the quality and composition of the aquatic andiparian communities. Thus, richness measurements do not weighhe importance of losing a particular species and fail to consider itsensitivity or importance in ecosystem functioning (i.e. all speciesave the same weight). For example, the loss of stenoic sensitiveaxa (Gutiérrez-Cánovas et al., 2013) or their substitution by gener-list or opportunistic species is a common occurrence as a responseo human disturbances (Devictor et al., 2008), which can resultn communities with similar richness but different composition.herefore, the use of holistic and integrative indicators considerings many river ecosystem components as possible (e.g. composi-ion, structure, functioning, diversity) or scoring the occurrencef species depending on its sensitiveness is preferable to partialnd/or simple parameters as richness measures.

.4. Management implications

The combined use of ecological condition indicators of bothacroinvertebrate and riparian vegetation provides a more real-

stic and complete picture of the status of river ecosystems leadingo improved planning, monitoring and evaluation of managementtrategies, restoration and conservation measures. Holistic andntegrative management seems the best option to reach the priorbjective of optimising river conservation as a whole (Europeanommission, 2000/60/EC).

Although both communities’ responses were similar, differentanagement measures could be proposed for each community

e.g. for strictly aquatic communities, a greater effort on mitigatingydromorphological alteration should be done). Our results suggestcale-dependent sensitivity of aquatic and riparian communitieso human pressures, which has general management implications.and use and hydrological planning at basin scale are essentialomplements to traditional, local and disconnected restorationeasures in order to successfully preserve and restore stream

cosystem integrity, biodiversity and longitudinal and lateral con-ectivity. Furthermore, reverting the human transformation of landse and limiting the current expansion of agriculture in the studyrea would help to minimise detriment to the river ecosystemsecause it is simultaneously one of the main causes of high wateremands and stream modification.

cknowledgements

We would like to thank to all of the members of the “Ecologíacuática” Research Group (UMU), especially J.A. Carbonell and F.icazo, as well as A. García-Ramos for their help and support withhe field and identification work. D.B. and D.S-F. were supportedy a predoctoral (FPU, AP2009-0432) and postdoctoral grant (Juane la Cierva program, JCI-2011-10529) respectively, both from the

panish Ministry of Economy and Competitiveness (MINECO) ofpain. We also want to thank to the Euromediterranean Institutef Water for the fieldwork funding received from the project

Hydrological classification of the rivers and streams in the Segura

ators 45 (2014) 456–464

basin and associated macroinvertebrate communities’. Finally, weare also very grateful to Cayetano Gutiérrez-Cánovas for his usefulcomments regarding the manuscript and Melissa Crim and JavierLloret for double-checking the English.

Appendix A. Supplementary data

Supplementary data associated with this article can befound, in the online version, at http://dx.doi.org/10.1016/j.ecolind.2014.04.051.

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