Multimetric assessment of macroinvertebrate responses to mitigation measures in a dammed and polluted river of Central Spain

Ecological Indicators 83 (2017) 356–367

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Original Articles

Multimetric assessment of macroinvertebrate responses to mitigation measures in a dammed and polluted river of Central Spain

MARK

Julio A. Camargo Unidad Docente de Ecología, Departamento de Ciencias de la Vida, Universidad de Alcalá, 28805 Alcalá de Henares (Madrid), Spain

A R T I C L E I N F O

A B S T R A C T

Keywords: Hydropower impoundment Industrial effluent Environmental impact mitigation Positive responses Benthic macroinvertebrates Multimetric assessment

In this research, structural and functional responses of benthic macroinvertebrates to mitigation measures (carried out in the dammed and polluted Duraton River, Central Spain, during the 1990s and 2000s) were examined by comparing physicochemical and biological data from the summer of 1987 with data from the summer of 2014. Mitigation measures resulted in significant increases in dissolved oxygen concentrations, as well as in significant reductions of fluoride (F−) pollution and short-term flow fluctuations. The macrobenthic community responded positively to improvements in river environmental conditions, exhibiting significant increases in abundance (total density, total biomass and EPT density) and diversity (total family richness and EPT richness) at impacted sampling sites. Furthermore, the presence of relatively sensitive benthic macroinvertebrates after mitigation measures (as indicated by increased values of BMWQ biotic indices) also was the main cause for observed reductions in the environmental impact caused by disturbance points (as indicated by decreased values of the EI index), and for the observed recovering of the trophic structure of the macrobenthic community, with macroinvertebrate scrapers as the functional feeding group most favored. These macroinvertebrate responses to mitigation measures were more marked at sampling sites that initially were more impacted (i.e., nearest to disturbance points), and less apparent at the sampling site that initially was less impacted (i.e., farthest to disturbance points). Within the hydropsychid assemblage, improvements in river environmental conditions clearly favored the presence of Hydropsyche pellucidula and Cheumatopsyche lepida at the expense of the other hydropsychid species. In spite of all monitored environmental improvements and macroinvertebrate positive responses, the need for additional mitigation measures was evident, particularly to reduce high turbidity levels and sedimentation of fine inorganic matter negatively affecting benthic macroinvertebrates downstream from the industrial effluent. Overall, it is concluded that the multimetric approach is an effective technique to assess macroinvertebrate responses to mitigation measures in river ecosystems.

1. Introduction There seems little doubt that man’s activities are seriously threatening the ecological integrity of most terrestrial and aquatic ecosystems on Earth. Within this tragic and real scenario, water pollution and its ensuing habitat degradation are among the most important anthropogenic causes of global change in river ecosystems. For example, surface runoff (from agriculture, primarily), wastewaters from livestock farming and inland aquaculture, and municipal sewage discharges (including effluents from sewage treatment plants without performing tertiary treatments) have significantly increased natural N and P fluxes into freshwater ecosystems around the world, this resulting in the widespread cultural eutrophication of surface waters (Harper, 1992; Smith et al., 1999; Räike et al., 2003; Camargo and Alonso, 2006; Schindler and Vallentyne, 2008; Ansari et al., 2011; Blaas and Kroeze, 2016). Likewise, excessive deposition of fine inorganic matter

E-mail address: [email protected]. http://dx.doi.org/10.1016/j.ecolind.2017.08.027 Received 6 March 2017; Received in revised form 17 July 2017; Accepted 9 August 2017 1470-160X/ © 2017 Elsevier Ltd. All rights reserved.

(particularly derived from agricultural catchments, road and channel constructions, mining operations, gravel pits, and industrial effluents) is becoming an anthropogenic environmental problem of increasing concern for river ecosystems worldwide (Wood and Armitage, 1997; Henley et al., 2000; Walling and Fang, 2003; Bilotta and Brazier, 2008; Bryce et al., 2010; Jones et al., 2012; Chapman et al., 2014; Naden et al., 2016). Additionally, numerous water chemical pollutants (e.g., acidification, cyanide, fluoride, detergents, heavy metals, insecticides and herbicides, PCBs, petroleum, pharmaceuticals, salinization) can cause significant adverse effects on aquatic organisms, this resulting in impaired populations and communities of freshwater ecosystems (Green and Trett, 1989; Rand, 1995; Camargo, 2003; Camargo and Alonso, 2006; Brain et al., 2008; Debenest et al., 2010; CañedoArgüelles et al., 2013; Stehle and Schulz, 2015; Berger et al., 2016; Della Rossa et al., 2017). The construction and operation of dams represent other important

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Fig. 1. General diagram of the study area in the middle Duraton River (Duero River Basin), showing the location of Burgomillodo Reservoir, the industrial plant, and sampling sites (D-1, D-2, D-3 and D-4).

The use of biological methods to assess water pollution and habitat degradation in river ecosystems has been traditionally recognized as an important complementary technique to conventional physicochemical surveys. This biomonitoring is usually based on the numerical value of several metrics and indices that integrate the ecological responses of aquatic organisms to environmental conditions, benthic macroinvertebrates being often regarded as the best indicators (Cairns and Dickson, 1973; Armitage et al., 1983; Washington, 1984; Hellawell, 1986; Extence et al., 1987; Rosenberg and Resh, 1993; Camargo, 1994; Hauer and Lamberti, 1996; Camargo et al., 2004; Bonada et al., 2006; Ziglio et al., 2006; Odume et al., 2012; Turley et al., 2016). In addition, a multimetric approach for practical biomonitoring with benthic macroinvertebrates is recommended since it can provide an integrated analysis of structural and functional attributes of the macrobenthic community, also pondering tolerances/sensitivities of benthic macroinvertebrates to water pollution and habitat degradation (Rosenberg and Resh, 1993; Camargo, 1994; Fore et al., 1996; Barbour and Yoder, 2000; Klemm et al., 2002; Camargo et al., 2004; Bonada et al., 2006; Ziglio et al., 2006; Odume et al., 2012; Seidel and Lüderitz, 2015; Serrano Balderas et al., 2016). In this investigation, a multimetric assessment of structural and functional responses of benthic macroinvertebrates to mitigation measures in the dammed and polluted Duraton River (Central Spain) was carried out by comparing physicochemical and biological data from the summer of 1987 (Camargo, 1989) with data from the summer of 2014. More specifically, four different categories of metrics and indices were estimated: (1) metrics of taxonomic composition and richness (including an index of environmental impact); (2) metrics of abundance (density and biomass); (3) metrics of trophic structure (contributions of functional feeding groups); and (4) metrics of environmental quality (biotic indices).

anthropogenic cause of global change in river ecosystems. At the end of the twentieth century, there were about 40,000 large dams (> 15 m in height) and more than 800,000 smaller dams worldwide, with nearly 75% of the world’s large rivers being fragmented by dams (Bednarek, 2001; Nilsson et al., 2005). In addition to river fragmentation, dam construction and operation can cause important adverse effects on the abundance and diversity of fluvial communities by modifying physicochemical conditions along river ecosystems. For example, reductions in river flow and matter transport, changes in water temperature and dissolved oxygen, and alterations in nutrient concentrations have been generally observed downstream from impoundments (Ward and Stanford, 1979; Petts, 1984; Camargo and Voelz, 1998; Lessard and Hayes, 2003; Camargo et al., 2005; Beck et al., 2012; Benitez-Mora and Camargo, 2014). Further environmental problems can arise when hydroelectric power generation induces short-term flow fluctuations (with extreme hydropeaking) downstream from dams (Moog, 1993; Lauters et al., 1996; Céréghino et al., 2002; Rehn, 2009; Bruder et al., 2016; Hauer et al., 2017). To reverse the impacts caused by human activities on natural populations and communities of river ecosystems, several mitigation and rehabilitation measures (e.g., improvements in sewage and wastewater treatments, recoveries of riparian zones, implementation of constructed riffles and small waterfalls, provision of adequate fish passages, establishment of downstream ecological flows, removal of dams and other man-made barriers) have been accomplished in rivers and streams around the world, but in general with limited success in restoring the ecological integrity of these aquatic ecosystems (Adams et al., 2002; Bednarek and Hart, 2005; Casper et al., 2006; Walther and Whiles, 2008; Lüderitz et al., 2011; Lorenz et al., 2012; Mueller et al., 2014; Kail et al., 2015; Muhar et al., 2016; Pan et al., 2016). Furthermore, field studies indicate that a deficient restoration of complex environmental factors, such as water quality and microhabitat heterogeneity, would be the primary cause for observed poor responses of benthic macroinvertebrates to mitigation and rehabilitation measures (Palmer et al., 2010; Verberk et al., 2010; Louhi et al., 2011; Sundermann et al., 2011; Ernst et al., 2012; Verdonschot et al., 2016). 357

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2. Materials and methods

Table 1 Geographic coordinates and environmental characteristics of sampling sites (D-1, D-2, D3 and D-4) in the study area.

2.1. Study area, mitigation measures, and sampling sites The Duraton River is located in Central Spain within the Duero River Basin (Fig. 1). It arises in “Peña Cebollera, Sistema Central” (> 2100 m a.s.l.) and flows for 106 km draining a basin of 1485 km2 before entering the Duero River. Field studies were conducted in the middle Duraton River (Segovia Province), in the vicinity of Burgomillodo Reservoir (Fig. 1), a eutrophic deep-release hydropower impoundment with a total capacity of 15 hm3, a surface area of 132 ha, and a maximum water depth of 40 m. Burgomillodo Dam is solely used for hydropower generation by discharging hypolimnial waters through three turbines with an installed power generation capacity of 3.83 MW. This gravity dam started to work in 1929, being heightened with an additional 15 m in 1953, currently exhibiting a structural height of 44 m and a crest length of 114 m. About 300 m downstream from Burgomillodo Dam, an industrial effluent enters Duraton River (Fig. 1). During the industrial process, hydrofluoric acid (HF) is used as a necessary, key reagent to induce charge separation of silica (SiO2) and feldspar from sandy materials, which are subsequently used as raw materials for manufacturing ceramics and glass at other industrial plants. As a consequence, high amounts of fluoride (F−), as well as suspended inorganic matter, are discharged by the industrial effluent into the Duraton River. Over the 27-year span between 1987 and 2014, three important mitigation measures were carried out in the middle Duraton River. On the one hand, the industrial wastewater treatment system was improved with better lime and limestone reactors to retain more efficiently fluoride ions as insoluble calcium fluoride (CaF2). Although other procedures may be used to remove fluoride ions from industrial waste waters (e.g., adsorption, ion exchange, reverse osmosis), chemical precipitation by using calcium containing materials is a relatively low-cost and high-efficiency method (Nath and Dutta, 2010; Waghmare and Arfin, 2015). On the other hand, a small man-made weir (about 1 m in height) was built 150 m downstream from Burgomillodo Dam to reinforce the re-oxygenation of hypolimnial waters released by the dam. It is worth noting that, in the summer of 2014, the impounded area between Burgomillodo Dam and this small man-made weir exhibited great abundance of submerged macrophytes (Potamogeton pectinatus and Myriophyllum spicatum, primarily), which clearly was an important factor for the re-oxygenation of hypolimnial waters via photosynthesis (the mean concentration of dissolved oxygen indeed was 8.7 ± 0.3 mg/L). Besides, short-term flow fluctuations were diminished as a result of more regular daily discharges from the dam, since hydropower production was much less dependent on the local immediate energy demand, and with the operating control of Burgomillodo Dam being moved to Bolarque Integrated Control Center (Guadalajara Province) in charge of controlling the hydropower production of several dams located in Central Spain. Actually, in 1987, the daily discharge of hypolimnial waters could fluctuated from 0.35 to 10.5 m3/s (Camargo, 1989) whereas a more regular daily discharge from the dam was apparent in 2014, being able to fluctuate between 1.2 and 2.9 m3/s (own estimates on the basis of river depth, river width and river current velocity). Increasing minimal discharges and decreasing maximal discharges is an important operational measure to reduce hydropeaking effects on river ecosystems (Bruder et al., 2016; Hauer et al., 2017). Four sampling sites were selected along the study area (Fig. 1). Geographic coordinates and environmental characteristics of sampling sites (D-1, D-2, D-3 and D-4) are presented in Table 1. D-1 was placed upstream from Burgomillodo Dam, within the limits of Natural Park “Hoces Río Duratón”. This sampling site was used as a reference station. D-2 was placed about 0.2 km downstream from Burgomillodo Dam, subsequent to the small man-made weir built to reinforce the re-oxygenation of hypolimnial waters released by the dam. D-3 and D-4 were

Latitude Longitude Elevation (m a.s.l.) River channel width (m) Predominant benthic substratum

D-1

D-2

D-3

D-4

41°17′34”N 3°50′40”W 892 6.4 Stony

41°20′21”N 3°53′15”W 839 12.1 Stony

41°20′24”N 3°53′19”W 838 8.2 Sandy/Stony

41°21′13”N 3°53′28”W 835 11.7 Stony

placed about 0.1 and 2.2 km downstream from the industrial effluent, respectively. The river bottom was mainly stony with cobbles and pebbles, but with an apparent deposition of fine inorganic matter at polluted sampling sites, particularly at D-3 where it was covered by a relatively thick layer of sandy (and silty) sediment. This deposition of fine inorganic matter was evident not only in 1987 (before mitigation measures), but also in 2014 (after mitigation measures). 2.2. Duraton River’s flow regime The natural flow regime of Duraton River (upstream from Burgomillodo Reservoir) is characterized by maximum flows during the winter (mainly in February) and minimum flows at the end of the summer (Fig. 2; Ministerio de Agricultura y Pesca, Alimentación y Medio Ambiente, 2017). However, downstream from Burgomillodo Reservoir, the Duraton River’s flow regime may be different, with maximum flows in March, April, May or even October, and minimum flows in June, November, December or even February (Fig. 2; Ministerio de Agricultura y Pesca, Alimentación y Medio Ambiente, 2017). Moreover, while maximum monthly river flows upstream from Burgomillodo Reservoir resulted to be similar when comparing the 1986–1987 and 2013–2014 hydrological years (Fig. 2), downstream from Burgomillodo Reservoir this comparison showed notable differences between both hydrological years, with maximum monthly river flows being much lower in 1986–87 than in 2013–2014 (Fig. 2). It hence seems evident that, upstream from Burgomillodo Reservoir, the precipitation pattern throughout the year is the key environmental factor determining the natural flow regime of Duraton River, whereas this natural flow regime may be significantly affected by dam operations downstream from Burgomillodo Reservoir. 2.3. Water sampling and analyses Sampling surveys to analyse water properties of the middle Duraton River were undertaken in the summer of 1987 and in the summer of 2014. Temperature, pH and dissolved oxygen were measured in situ at each sampling site on each sampling survey using specific meters in accordance with American Public Health Association (1980, 1998). Additionally, water samples were collected with clean polyethylene containers to analyze in the laboratory fluoride and turbidity. Fluoride was analysed by the standard TISAB potentiometric method (American Public Health Association, 1980) in the 1987 sampling survey, and by the standard SPADNS colorimetric method (American Public Health Association, 1998) in the 2014 sampling survey. Turbidity was measured only in the 2014 sampling survey using a standard turbidimeter (American Public Health Association, 1998). 2.4. Macroinvertebrate sampling and analyses Benthic macroinvertebrates were collected using a quantitative cylindrical bottom sampler (Hellawell, 1986; Hauer and Lamberti, 1996; Wetzel and Likens, 2000), which enclosed a sampling area of about 0.1 m2 and had a capturing net with a mesh size of 250 μm. Five riffle bottom samples were randomly taken at each sampling site on each 358

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Fig. 2. Mean monthly flows (m3/s) in the Duraton River, upstream and downstream from Burgomillodo Reservoir, during the 1986–1987 and 2013–2014 hydrological years. Upstream flows were measured at the 2012 gauging station (3°44′11”W, 41°18′07”N) and downstream flows were measured at the 2161 gauging station (3°57′58”W, 41°26′07”N) (source of data: Ministerio de Agricultura y Pesca, Alimentación y Medio Ambiente, 2017).

with benthic macroinvertebrates (Rosenberg and Resh, 1993; Fore et al., 1996; Barbour and Yoder, 2000; Klemm et al., 2002; Camargo et al., 2004; Bonada et al., 2006; Ziglio et al., 2006; Odume et al., 2012; Seidel and Lüderitz, 2015; Serrano Balderas et al., 2016). To estimate the environmental impact caused by disturbance points on the structure (family richness and family composition) of the macroinvertebrate community at each impacted sampling site on each sampling survey, Camargo’s (1990) index of environmental impact was calculated: EI (%) = (2A − B − C) × 50/A. Where A is the total taxonomic richness (total family richness in this study) at the reference station, B is the total taxonomic richness at an impacted sampling site (D-2, D-3 or D-4 in this study), and C is the taxonomic richness common to both the reference station and the impacted sampling site. This index usually takes percentage values between 0 (no environmental impact) and 100 (maximum environmental impact), but it can exhibit negative values when environmental disturbances cause some positive effects on the structure of disturbed communities (as in the case of the intermediate disturbance hypothesis, one of the most influential theories in ecology). To examine changes in the trophic structure of the macroinvertebrate community at each sampling site on each sampling survey, benthic macroinvertebrates were allocated to five functional feeding groups in accordance with Tachet et al. (2003), Merritt et al. (2008) and Thorp and Covich (2010): shredders basically feed on coarse particulate organic matter (e.g., Gammaridae amphipods, Capniidae, Leuctridae and Nemouridae plecopterans, Limnephilidae trichopterans, Tipulidae dipterans); scrapers mainly feed on periphyton and perilithon (e.g., Ancylidae and Physidae gastropods, Elmidae coleopterans, Baetidae and Heptageniidae ephemeropterans, Glossosomatidae, Hydroptilidae and Psychomyiidae trichopterans, Blephariceridae dipterans); collector-gatherers largely feed on fine organic detritus (e.g., Naididae oligochaetes, Caenidae, Ephemerellidae and Potamanthidae ephemeropterans,

sampling survey. All macroinvertebrate samples were preserved in 4% formalin until laboratory analyses. In the laboratory, benthic macroinvertebrates were identified and counted using a light binocular stereomicroscope. Taxonomic identification was basically carried out to the family level following Tachet et al. (1981, 2003). This family level has been recommended for practical biomonitoring of freshwater pollution and habitat degradation with benthic macroinvertebrates (Armitage et al., 1983; Hellawell, 1986; Extence et al., 1987; Rosenberg and Resh, 1993; Camargo et al., 2004; Ziglio et al., 2006; Odume et al., 2012). Additionally, some macroinvertebrate taxa, particularly netspinning caddisfly larvae of the Hydropsychidae family, were identified to the species level following Consiglio (1980), Edington and Hildrew (1981), Belfiore (1983), Moretti (1983), Fernández-Lop (1987), Camargo and García de Jalón (1992), and Zamora-Muñoz et al. (1995). After identification and counting, macroinvertebrate samples were dried in an oven at 60 °C for 24 h in order to estimate total biomass (dry-weight). 2.5. Macroinvertebrate metrics and indices Metrics of abundance and taxonomic richness were estimated for the whole macroinvertebrate community at each sampling site on each sampling survey. Macroinvertebrate abundance is expressed as total density (total number of individuals) per square meter of benthic substratum, and as total biomass (total dry-weight) per square meter of benthic substratum. Macroinvertebrate richness is expressed as the total number of taxonomic families (total family richness) per sampling unit. In addition, EPT density (number of individuals in Ephemeroptera, Plecoptera and Trichoptera) and EPT richness (number of taxonomic families in Ephemeroptera, Plecoptera and Trichoptera) were calculated. All these metrics of abundance and taxonomic richness are generally used within the multimetric approach for practical biomonitoring 359

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the use of a-BMWQ biotic index has been recommended for assessing environmental quality in relation to the tolerance/sensitivity of benthic macroinvertebrates to freshwater pollution and habitat degradation (Camargo, 1993, 1994). In this investigation, both biotic indices (tBMWQ and a-BMWQ) were calculated, being expressed as their respective values per sampling unit.

Table 2 Mean (n = 4–17) values of water physicochemical parameters at sampling sites (D-1, D-2, D-3 and D-4) in the 1987 and 2014 sampling surveys. Significant (P < 0.05) differences between the reference station (D-1) and impacted/polluted sampling sites (D-2, D-3 and D-4) on each sampling survey are indicated with the letter a. Significant (P < 0.05) differences between sampling surveys at each sampling site are indicated with the letter b. ± SD between parentheses. D-1

Water temperature (°C) pH Dissolved oxygen (mg O2/L) Fluoride (mg F−/L) Turbidity (NTU)

D-2

D-3

D-4

2.6. Statistical analyses

1987

2014

1987

2014

1987

2014

1987

2014

16.2 (0.9)

16.9 (0.7)

15.7b (0.4)

16.5b (0.3)

15.9 (0.5)

16.7 (0.4)

16.3 (0.8)

17.1 (0.6)

8.1 (0.0) 8.7 (0.6)

8.2 (0.1) 9.1 (0.5)

7.6a,b (0.1) 4.0a,b (0.7)

8.0b (0.2) 8.9b (0.5)

7.7a,b (0.1) 5.8a,b (1.1)

8.2b (0.2) 9.0b (0.6)

7.8a (0.1) 6.7a,b (1.2)

8.1 (0.3) 9.3b (0.8)

0.1 (0.0) –

0.1 (0.0) 4.6 (1.7)

0.1 (0.0) –

0.1 (0.0) 3.3 (0.9)

6.8a,b (5.7) –

0.6a,b (0.2) 55.2a (11.8)

2.7a,b (2.1) –

0.4a,b (0.2) 20.5a (6.1)

Since parametric methods usually have more statistical power than nonparametric methods for detecting that the null hypothesis is false (Sokal and Rohlf, 1995), I have preferred to use a parametric method in order to reduce the type II error (i.e., failing to detect physicochemical and biological changes at impacted/polluted sampling sites, and failing to detect positive responses of benthic macroinvertebrates to mitigation measures). Accordingly, t-tests were performed to check significant (P < 0.05) differences between pair of independent means: on the one hand, I compared mean values of water physicochemical parameters and biological metrics and indices at the reference station (D-1) with mean values at each impacted/polluted sampling site (D-2, D-3 or D-4); on the other hand, I compared mean values of water physicochemical parameters and biological metrics and indices in the 1987 sampling survey with mean values in the 2014 sampling survey. Normality and homoscedasticity of data were assumed for physicochemical parameters and biological metrics and indices (Sokal and Rohlf, 1995). Statistical analyses were performed using the SPSS software.

Chironomidae dipterans); collector-filterers primarily feed on organic material suspended in the water column (e.g., bivalves, Oligoneuriidae ephemeropterans, Brachycentridae, Hydropsychidae and Philopotamidae trichopterans, Simuliidae dipterans); and predators feed on animal preys (e.g., planarians, leeches, damselflies and dragonflies, Perlidae plecopterans, Gyrinidae coleopterans, Rhyacophilidae and Polycentropodidae trichopterans, Anthomyiidae, Athericidae and Empididae dipterans). In this investigation, relative contributions (%) of functional feeding groups to the macroinvertebrate community were calculated on the basis of density estimates. Lastly, to assess environmental quality in relation to the tolerance/ sensitivity of benthic macroinvertebrates to freshwater pollution and habitat degradation at each sampling site on each sampling survey, the Biological Monitoring Water Quality (BMWQ) score system (Camargo, 1993, 1994) was applied. The BMWQ score system is based on the relative tolerance of Iberian macroinvertebrate families to freshwater pollution and habitat degradation (particularly to organic pollution and nutrient enrichment), following the family-level operating structure of the Biological Monitoring Working Party score system for biomonitoring river environmental quality in Great Britain (National Water Council, 1981; Armitage et al., 1983). According to Camargo (1993, 1994), pollution sensitive families can exhibit high scores between 11 and 15 (e.g., Planariidae, Gammaridae, Perlidae, Capniidae, Leuctridae, Nemouridae, Ephemerellidae, Heptageniidae, Oligoneuriidae, Libellulidae, Limnephilidae, Brachycentridae, Glossosomatidae, Philopotamidae, Sericostomatidae, Psychomyiidae, Rhyacophilidae, Elmidae, Athericidae, Blephariceridae); pollution tolerant families can exhibit low scores between 1 and 5: (e.g., Naididae, Glossiphoniidae, Physidae, Chironomidae, Syrphidae); and families with intermediate tolerances can exhibit middle scores between 6 and 10 (e.g., Dugesiidae, Erpobdellidae, Ancylidae, Baetidae, Caenidae, Potamanthidae, Hydropsychidae, Hydroptilidae, Gyrinidae, Simuliidae, Tipulidae, Anthomyiidae, Empididae). After macroinvertebrate samples have been analysed, the totalBMWQ (t-BMWQ) biotic index and the average-BMWQ (a-BMWQ) biotic index may be calculated (Camargo, 1993, 1994): the t-BMWQ biotic index is estimated by summing the individual scores of all families present in a sample, being able to take values between 1 (very poor environmental quality) and more than 270 (excellent environmental quality); the a-BMWQ biotic index is estimated by dividing the value of t-BMWQ index by the number of families in the sample, being able to take values between 1 (very poor environmental quality) and more than 12 (excellent environmental quality). Because a-BMWQ values are less sensitive to sample size (sampling effort), family richness (macrobenthic diversity) and seasonal changes than t-BMWQ values,

3. Results Mean values of water physicochemical parameters at each sampling site for the 1987 and 2014 sampling surveys are presented in Table 2. In 1987, the daily discharge of hypolimnial waters from Burgomillodo Dam caused a significant (P < 0.05) deficit of dissolved oxygen and decreased pH values at downstream sampling sites (D-2, D-3 and D-4), and with the industrial effluent also causing high fluoride pollution at polluted sampling sites (D-3 and D-4) (Table 2). Conversely, in 2014, the daily discharge of hypolimnial waters from Burgomillodo Dam did not cause a deficit of dissolved oxygen and decreased pH values downstream from the dam, with dissolved oxygen concentrations and pH values being significantly (P < 0.05) higher than those in 1987 (Table 2). Moreover, the industrial effluent caused low fluoride pollution downstream, with fluoride concentrations at D-3 and D-4 being significantly (P < 0.05) much lower than those in 1987 (Table 2). However, the industrial effluent caused significant (P < 0.05) increases in turbidity at D-3 and D-4 (Table 2). Relative abundances of macroinvertebrate families at each sampling site for the 1987 and 2014 sampling surveys are presented in Table 3. In both sampling surveys, Gammaridae (Echinogammarus calvus) amphipods, and Brachycentridae (Brachycentrus subnubilus), Glossosomatidae (Agapetus laniger) and Limnephilidae trichopterans were collected at the reference station (D-1), but not at impacted sampling sites (D-2, D-3 and D-4). Although the family composition of the benthic macroinvertebrate community at D-1 in the 2014 sampling survey was similar to that in the 1987 sampling survey, some differences were evident (Table 3): while Nemouridae (Protonemura meyeri) plecopterans, Oligoneuriidae (Oligoneuriella rhenana) ephemeropterans, Gyrinidae coleopterans, and Tipulidae and Empididae dipterans were not found, Leuctridae plecopterans and Blephariceridae dipterans were collected. All in all, the most important changes between 1987 and 2014 sampling surveys were observed at D-2 and D-3 sampling sites. In the 1987 sampling survey, Erpobdellidae leeches, Chironomidae dipterans, and Ephemerellidae (Ephemerella ignita), Caenidae (Caenis moesta) and Baetidae ephemeropterans were significantly (P < 0.05) the predominant benthic macroinvertebrates at D-2, and with only three families (Chironomidae, Erpobdellidae and Baetidae) being present at D-3 (Table 3). In contrast, in the 2014 sampling survey, Dugesiidae planarians, 360

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and H. lobata (Fig. 3). Although, in the 2014 sampling survey, the most abundant hydropsychid species at D-1 still was H. pellucidula, being followed in importance by Ch. lepida, H. siltalai, H. bulbifera and H. exocellata (H. lobata was not found), and with no net-spinning caddisfly larva being found at D-3, significant differences were observed at D-2 and D-4 with regard to the findings of 1987 (Fig. 3). At D-2, not only the abundance of net-spinning caddisfly larvae had significantly (P < 0.05) changed, but also the number and predominance of hydropsychid species, with H. pellucidula as the most abundant species, and being followed in importance by Ch. lepida and H. exocellata (Fig. 3). At D-4, Ch. lepida was the most abundant hydropsychid species (P < 0.05), being followed in importance by H. pellucidula, H. exocellata, H. bulbifera and H. siltalai (Fig. 3). Mean values of macroinvertebrate metrics and indices at each sampling site for the 1987 and 2014 sampling surveys are presented in Table 4. In both sampling surveys, the highest mean values of total family richness (22.6 and 20.2), total density (4948 and 4026), total biomass (8458 and 5642), EPT richness (12.4 and 12.0), EPT density (2990 and 1872), and t-BMWQ (224 and 203) and a-BMWQ (9.9 and 10.1) indices were significantly (P < 0.05) estimated at D-1 (the reference station), whereas the lowest mean values were estimated at D-3 (the most impacted sampling site), with D-2 and D-4 exhibiting in-between mean values (Table 4). Nevertheless, significant differences between 1987 and 2014 sampling surveys were evident, particularly regarding D-2 and D-3 sampling sites (Table 4). Indeed, mean values of macroinvertebrate metrics and indices at D-2 and D-3 were significantly (P < 0.05) higher in the 2014 sampling survey than in the 1987 sampling survey: for example, at D-3, mean values of total family richness (4.8), total density (514), EPT richness (2.6), EPT density (184), and t-BMWQ (36.8) and a-BMWQ (7.7) indices in 2014, compared to mean values of total family richness (1.4), total density (116), EPT richness (0.2), EPT density (2), and t-BMWQ (6.4) and a-BMWQ (3.4) indices in 1987 (Table 4). In the case of the farthest sampling site (D-4), differences between sampling surveys were less apparent. Relative contributions (%) of functional feeding groups to the trophic structure of the macroinvertebrate community at each sampling site for the 1987 and 2014 sampling surveys are shown in Fig. 4. In 1987, collector-filterers were the predominant functional feeding group at D-1 (51.63), mainly because of Oligoneuriidae ephemeropterans, whereas collector-gatherers were the predominant functional feeding group downstream from Burgomillodo Reservoir, particularly at D-2 (68.71) and D-3 (86.21) (Fig. 4). The following functional feeding groups in importance were predators at D-2 (14.62) and D-3 (12.07), and collector-filterers at D-4 (36.89), but with this functional feeding group being absent at D-3 and poorly represented at D-2 (4.96) (Fig. 4). Mean contributions of scrapers and shredders were relatively low at all sampling sites, particularly those of shredders which were absent at D-3 (Fig. 4). Nevertheless, in 2014, despite collector-gatherers were the predominant functional feeding group at D-1 (29.60), D-3 (64.59) and D-4 (34.50), scrapers showed significant (P < 0.05) increases in their relative contributions, particularly at D-2 (37.60) and D-3 (30.74) where they were the first and second functional feeding group in importance, respectively (Fig. 4). Similarly, shredders increased significantly (P < 0.05) their relative contributions at D-1 (14.53), D-2 (2.82) and D-4 (4.56), but still they were absent at D-3 (Fig. 4). Conversely, mean contributions of predators decreased at sampling sites with regard to their values in the 1987 sampling survey (Fig. 4), particularly at D-2 (4.99) and D-3 (4.67). Collector-filterers increased significantly (P < 0.05) at D-2 (17.86), but this functional feeding group decreased at D-1 (25.64) and D-4 (31.60), and still it was absent at D-3 (Fig. 4). Estimated values of the environmental impact (%) caused by disturbance points on the structure (family richness and family composition) of the benthic macroinvertebrate community at disturbed sampling sites (D-2, D-3 and D-4) before (1987) and after (2014) mitigation measures are shown in Fig. 5. In the 1987 sampling survey, the highest

Table 3 Relative abundances (%) of macroinvertebrate families at sampling sites (D-1, D-2, D-3 and D-4) in the 1987 and 2014 sampling surveys. D-1

Dugesiidae Erpobdellidae Naididae Ancylidae Physidae Gammaridae Baetidae Oligoneuriidae Heptageniidae Ephemerellidae Caenidae Potamanthidae Capniidae Leuctridae Nemouridae Elmidae Gyrinidae Glossosomatidae Rhyacophilidae Hydroptilidae Hydropsychidae Polycentropodidae Psychomyiidae Limnephilidae Brachycentridae Philopotamidae Tipulidae Blephariceridae Simuliidae Chironomidae Empididae Anthomyiidae

D-2

D-3

D-4

1987

2014

1987

2014

1987

2014

1987

2014

1.70 0.00 0.12 1.66 0.00 2.22 10.43 21.95 2.87 2.99 1.70 0.00 2.99 0.00 1.21 8.41 0.65 0.28 0.73 0.04 8.45 0.00 0.69 0.08 0.65 5.38 0.04 0.00 15.20 9.40 0.16 0.00

2.46 0.00 0.60 4.17 0.00 6.95 7.50 0.00 3.34 5.92 6.36 0.00 5.12 1.29 0.00 7.25 0.00 1.04 0.55 0.84 6.95 0.00 2.73 1.17 1.24 3.60 0.00 0.35 13.85 16.72 0.00 0.00

0.00 10.60 0.96 4.42 0.00 0.00 6.73 0.00 0.00 9.46 28.30 1.69 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.34 1.58 3.91 0.00 0.00 0.00 0.00 0.22 0.00 3.38 28.30 0.00 0.11

1.69 1.12 3.89 3.32 2.26 0.00 11.97 0.00 1.91 3.25 13.62 0.00 2.82 0.00 0.00 8.12 0.00 0.00 0.49 2.75 4.38 1.69 7.27 0.00 0.00 2.54 0.00 0.00 10.94 15.97 0.00 0.00

0.00 12.07 0.00 0.00 0.00 0.00 1.72 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 86.21 0.00 0.00

0.00 4.67 0.00 6.23 2.33 0.00 14.01 0.00 0.00 0.00 10.51 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 8.17 0.00 0.00 0.00 0.00 0.00 0.00 54.08 0.00 0.00

0.00 0.40 0.50 5.39 0.00 0.00 4.40 0.00 0.00 13.11 3.96 3.35 3.25 0.00 0.00 5.79 0.00 0.00 1.42 0.00 29.17 3.15 1.83 0.00 0.00 0.00 0.20 0.00 7.72 16.36 0.00 0.00

0.38 1.37 5.32 4.64 0.00 0.00 6.23 0.00 1.52 4.26 5.09 0.00 4.56 0.00 0.00 7.52 0.00 0.00 0.76 0.99 14.50 1.22 4.71 0.00 0.00 0.00 0.00 0.00 17.10 19.83 0.00 0.00

Heptageniidae ephemeropterans, Capniidae (Capnia bifrons) plecopterans, Philopotamidae (Chimarra marginata), Psychomyiidae (Psychomyia pussila) and Rhyacophilidae trichopterans, Elmidae coleopterans and Physidae (Physella acuta) snails were collected at D-2, and with seven families (Erpobdellidae, Chironomidae, Caenidae, Baetidae, Psychomyiidae, Ancylidae and Physidae) being present at D-3 (Table 3). However, in both sampling surveys, Chironomidae dipterans were significantly (P < 0.05) the most abundant macroinvertebrates at D-3. Regarding the farthest sampling site (D-4), the family composition of the macrobenthic community was relatively similar between 1987 and 2014 sampling surveys, but with some significant differences being observed (Table 3): while Dugesiidae planarians, Heptageniidae ephemeropterans, and Hydroptilidae trichopetrans were collected in 2014 but not in 1987, Potamanthidae (Potamanthus luteus) ephemeropterans and Tipulidae dipterans were found in 1987 but not in 2014. Furthermore, in the 1987 sampling survey, Hydropsychidae trichopterans were significantly (P < 0.05) the most abundant benthic macroinvertebrates at D-4, whereas in the 2014 sampling survey the most abundant macroinvertebrates were significantly (P < 0.05) Chironomidae dipterans. Mean densities (individuals/m2) of Hydropsychidae species at each sampling site on each sampling survey are shown in Fig. 3. In the 1987 sampling survey, the most abundant hydropsychid species at D-1 was significantly (P < 0.05) Hydropsyche pellucidula, being followed in importance by H. bulbifera, Cheumatopsyche lepida, H. exocellata, H. lobata and H. siltalai (Fig. 3). However, only two hydropsychid species (H. bulbifera and H. exocellata) were found at D-2 (both species exhibiting relatively low abundances), and no net-spinning caddisfly larva was found at D-3 (Fig. 3). In contrast, five hydropsychid species were found at D-4, with H. bulbifera as the most abundant species (P < 0.05), and being followed in importance by H. siltalai, H. exocellata, H. pellucidula 361

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Fig. 3. Mean (n = 5) densities of Hydropsychidae species at sampling sites (D-1, D-2, D-3 and D-4) in the 1987 and 2014 sampling surveys.

Table 4 Mean (n = 5) values of macroinvertebrate metrics and indices at sampling sites (D-1, D-2, D-3 and D-4) in the 1987 and 2014 sampling surveys. Significant (P < 0.05) differences between the reference station (D-1) and impacted/polluted sampling sites (D-2, D-3 and D-4) on each sampling survey are indicated with the letter a. Significant (P < 0.05) differences between sampling surveys at each sampling site are indicated with the letter b. ± SD between parentheses. D-1

Total family richness (# families/sampling unit) Total density (individuals/m2) Total biomass (mg dry-weight/m2) EPT richness (# families/sampling unit) EPT density (individuals/m2) t-BMWQ index (index value/sampling unit) a-BMWQ index (index value/sampling unit)

D-2

D-3

D-4

1987

2014

1987

2014

1987

2014

1987

2014

22.6b (1.2) 4948 (619) 8458 (2789) 12.4 (1.1) 2990 (826) 224 (13.1) 9.9 (0.1)

20.2b (1.1) 4026 (663) 5642 (1609) 12.0 (0.7) 1872 (316) 203 (12.3) 10.1 (0.1)

12.2a,b (0.8) 1902a,b (186) 2381a (544) 3.6a,b (0.9) 968a,b (239) 104a,b (8.1) 8.5a,b (0.1)

16.4a,b (1.1) 2834a,b (527) 3697 (1018) 8.8a,b (0.8) 1494b (265) 149a,b (10.9) 9.1a,b (0.1)

1.4a,b (1.2) 116a,b (71) 164a,b (98) 0.2a,b (0.4) 2a,b (4) 6.4a,b (6.8) 3.4a,b (2.2)

4.8a,b (0.9) 514a,b (168) 754a,b (203) 2.6a,b (0.5) 184a,b (61) 36.8a,b (7.5) 7.7a,b (0.5)

13.8a (1.3) 1968a,b (272) 4192a (827) 7.4a (1.1) 1252a (274) 123a,b (9.5) 8.9a,b (0.1)

15.0a (0.7) 2632a,b (349) 3470 (796) 7.6a (0.9) 1130a (198) 144a,b (10.3) 9.6a,b (0.2)

middle Duraton River. On the one hand, improvements in the industrial wastewater treatment system (by the use of better lime and limestone reactors to retain more efficiently fluoride ions as insoluble calcium fluoride) significantly reduced fluoride concentrations downstream from the industrial effluent (Table 2). In fact, fluoride concentrations measured downstream from the industrial effluent in the 2014 sampling survey were about the maximum safe level of 0.5 mg F−/L recommended for protecting sensitive freshwater fauna (World Health Organization, 2002; Camargo, 2003). Nevertheless, increased turbidity downstream from the industrial effluent was considerable (Table 2), with deposition of fine inorganic matter on the river bottom of polluted

environmental impact was estimated at D-3 (90), being followed in importance by D-2 (52) and D-4 (42) (Fig. 5). In the 2014 sampling survey, all estimated values of environmental impact were lower than their respective values in 1987, with D-3 still exhibiting the highest value (72.7), but being followed in importance by D-4 (27.3) and D-2 (20.5) (Fig. 5). 4. Discussion It is evident that mitigation measures carried out during the 1990s and 2000s have markedly improved environmental conditions of the 362

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Fig. 4. Relative contributions (%) of functional feeding groups to the macroinvertebrate community at sampling sites (D-1, D-2, D-3 and D-4) in the 1987 and 2014 sampling surveys.

Muhar et al., 2016). In this case study, however, submerged macrophytes, inhabiting the impounded area between the small man-made weir and Burgomillodo Dam, were clearly responsible via photosynthesis for the re-oxygenation of hypolimnial waters released by the dam (the mean concentration of dissolved oxygen at that impounded area was 8.7 ± 0.3 mg/L). In addition, short-term flow fluctuations were minimized as a result of more regular daily discharges from Burgomillodo Dam, being able to fluctuate between 0.35 and 10.5 m3/s in 1987 (Camargo 1989) and between 1.2 and 2.9 m3/s in 2014 (own estimates on the basis of river depth, river width and river current velocity). Extreme short-term flow

sampling sites, particularly at D-3 where it was covered by a relatively thick layer of sandy (and silty) sediment. Although turbidity was not measured in 1987, the high deposition of fine inorganic matter on the river bottom of D-3 already was evident. On the other hand, the construction of a small man-made weir, to reinforce the re-oxygenation of hypolimnial waters released by Burgomillodo Dam, significantly increased dissolved oxygen concentrations downstream from the dam (Table 2). Implementation of constructed riffles and small waterfalls is an important rehabilitating technique to recover and maintain water levels of dissolved oxygen in river ecosystems (Adams et al., 2002; Darby and Sear, 2008; Lüderitz et al., 2011; Roni and Beechie, 2013; 363

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Fig. 5. Estimated values of the environmental impact (%) caused by disturbance points on the structure (family richness and family composition) of the benthic macroinvertebrate community at disturbed sampling sites (D-2, D-3 and D-4) before (1987) and after (2014) mitigation measures in the middle Duraton River (Central Spain).

fluctuations is one of the most important environmental factors negatively affecting freshwater populations and communities downstream from hydropower impoundments (Moog, 1993; Lauters et al., 1996; Céréghino et al., 2002; Rehn, 2009; Bruder et al., 2016; Hauer et al., 2017), with increasing minimal discharges and decreasing maximal discharges as an important operational measure to reduce hydropeaking effects on river ecosystems (Bruder et al., 2016; Hauer et al., 2017). Responses of benthic macroinvertebrates to mitigation measures in the middle Duraton River may be considered clearly positive owing to significant increases in abundance (total density, total biomass and EPT density) and diversity (total family richness and EPT richness) of the macroinvertebrate community at D-2 and D-3 in the 2014 sampling survey (Table 4). The fact that these positive responses of benthic macroinvertebrates to mitigation measures were less apparent at D-4 is not surprising since this sampling site was located farthest to disturbance points (Burgomillodo Dam and the industrial effluent) and, therefore, it was originally less impacted than D-2 and D-3 (Table 4; Fig. 5). Several field studies have already showed that positive responses of aquatic populations and communities to mitigation and rehabilitation measures in freshwater ecosystems can occur more rapidly and markedly at areas that initially were more impacted (Adams et al., 2002; Lüderitz et al., 2011; Lorenz et al., 2012; Kail et al., 2015; Muhar et al., 2016). The positive responses of benthic macroinvertebrates to mitigation measures are also confirmed by the lower values of environmental impact estimated at disturbed sampling sites in the 2014 sampling survey (Fig. 5). This reduction of environmental impact was more marked at D-2 than at D-3 as a probable consequence of the increased turbidity downstream from the industrial effluent, with an apparent deposition of fine inorganic matter on the river bottom of D-3. Field studies conducted in different rivers and streams have already showed that increases in water turbidity can reduce light transmission and intensify siltation of suspended solids, these altered processes causing reductions in the abundance of sensitive benthic invertebrates owing to decreases in food supply and quality, and also because of alterations in available and suitable benthic substratum (Wood and Armitage, 1997; Henley et al., 2000; Bilotta and Brazier, 2008; Bryce et al., 2010; Jones et al., 2012; Chapman et al., 2014; Naden et al., 2016). Significant increases in abundance and diversity of the macroinvertebrate community, as well as decreases in the environmental impact caused by disturbance points, were mainly due to the presence of relatively sensitive benthic macroinvertebrates at impacted sampling sites after mitigation measures (in the 2014 sampling survey), all of which however were absent (or exhibited much lower abundances) before mitigation measures (in the 1987 sampling survey). This was the case for macroinvertebrate families such as Dugesiidae, Ancylidae, Simuliidae, Philopotamidae, Psychomyiidae, Hydropsychidae,

Rhyacophilidae, Elmidae, Capniidae and Heptageniidae at D-2, and Baetidae, Caenidae, Psychomyiidae and Ancylidae at D-3 (Table 3). The presence of benthic macroinvertebrates more sensitive (or less tolerant) to water pollution and habitat degradation also was evident because mean values of BMWQ biotic indices at disturbed sampling sites were significantly higher in the 2014 sampling survey than in the 1987 sampling survey (Table 4). Nevertheless, in 2014, mean values of macroinvertebrate metrics and indices still were much lower at D-3 than at D-2 and D-4 (Table 4), as a probable consequence of higher turbidity levels and deposition of fine inorganic matter at D-3. Regarding the hydropsychid assemblage (Hydropsychidae family), significant changes in species abundance and richness at D-2 may be reasonably explained because of improvements in river environmental conditions. The fact that, in the 1987 sampling survey, only Hydropsyche bulbifera and H. exocellata were found at D-2, both species with relatively low abundances (Fig. 3), whereas in the 2014 sampling survey the species abundance and richness of the hydropsychid assemblage had significantly increased, with H. pellucidula and Cheumatopsyche lepida as the most abundant species, being followed in importance by H. exocellata (Fig. 3), it was probably due to significant increases in dissolved oxygen concentrations and reductions of shortterm flow fluctuations downstream from Burgomillodo Dam. Laboratory and field studies indicate that H. pellucidula might be a relatively sensitive species to low levels of dissolved oxygen, and also to extreme short-term flow fluctuations, this species tending to be replaced by other more tolerant Hydropsyche species in disturbed river ecosystems (Becker, 1987; Boon, 1987; Petts, 1984; Moog, 1993; Engels et al., 1996; Englund et al., 1997). Similarly, as a likely consequence of significant increases in dissolved oxygen concentrations and reductions of short-term flow fluctuations downstream from Burgomillodo Dam, H. pellucidula and Ch. lepida also were the most abundant hydropsychid species at the farthest sampling site (D-4). Relatively sensitive benthic macroinvertebrates collected along the study area in the 2014 sampling survey would be responsible not only for significant increases in abundance and diversity of the macroinvertebrate community, but also for significant changes in its trophic structure (Fig. 4). Actually, the presence of scrapers such as Ancylidae, Elmidae, Heptageniidae and Psychomyiidae, as well as the presence of Capniidae shredders and Simuliidae and Hydropsychidae collector-filterers, resulted in a more complex trophic structure of the macroinvertebrate community at D-2 in the 2014 sampling survey, and with this trophic structure becoming less dissimilar to that at the reference station (Fig. 4). Similarly, the presence of Psychomyiidae, Ancylidae and Physidae at D-3 resulted in higher abundance of scrapers just below the industrial effluent in the 2014 sampling survey (Fig. 4). Regarding the farthest sampling site (D-4), changes in the trophic structure of the macroinvertebrate community were less apparent, with collectorgatherers and collector-filterers being the most abundant functional 364

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Acknowledgements

feeding groups in the 1987 and 2014 sampling surveys (Fig. 4). Although the use of simple metrics or indices to assess ecosystem health and integrity has been criticized because potentially important ecological information may be lost (Karr, 1991; Suter, 1993; Scrimgeour and Wicklum, 1996; Rapport et al., 1998), supporters of the multimetric approach with benthic macroinvertebrates have argued that this technique may be enough robust for practical biomonitoring since it provides an integrated analysis of structural and functional attributes of the macrobenthic community, also pondering tolerances/ sensitivities of benthic macroinvertebrates to water pollution and habitat degradation (Rosenberg and Resh, 1993; Camargo, 1994; Fore et al., 1996; Barbour and Yoder, 2000; Klemm et al., 2002; Camargo et al., 2004; Bonada et al., 2006; Ziglio et al., 2006; Odume et al., 2012; Seidel and Lüderitz, 2015; Serrano Balderas et al., 2016). According to the obtained results in the dammed and polluted Duraton River, it seems evident that the multimetric approach with benthic macroinvertebrates has resulted to be a sensitive and suitable technique for biomonitoring the success of mitigation measures, properly differentiating macroinvertebrate responses between sampling surveys (before and after mitigation measures) and between sampling sites along the study area. In this regard, it is worth noting that Seidel and Lüderitz (2015) have recently reported that the multimetric approach was more effective for assessing river restoration success than the evaluation methods according to the European Water Framework Directive (WFD, 2000/60/EC).

Funds for this research came from the Spanish Ministry of Science and Innovation (Research project CGL2011-28585). The University of Alcala provided logistical support for carrying out field and laboratory studies. Valuable comments and suggestions of two anonymous reviewers are also acknowledged. References Adams, S.M., Hill, W.R., Peterson, M.J., Ryon, M.G., Smith, J.G., Stewart, A.J., 2002. Assessing recovery in a stream ecosystem: applying multiple chemical and biological endpoints. Ecol. Appl. 12, 1510–1527. American Public Health Association, 1980. Standard Methods for the Examination of Water and Wastewater, 15th ed. APHA-AWWA-WPCF, Washington, DC. American Public Health Association, 1998. Standard Methods for the Examination of Water and Wastewater, 20th ed. APHA-AWWA-WPCF, Washington, DC. Ansari, A.A., Gill, S.S., Lanza, G.R., Rast, W. (Eds.), 2011. Eutrophication: Causes, Consequences and Control. Springer, Dordrecht. Armitage, P.D., Moss, D., Wright, J.F., Furse, M.T., 1983. The performance of a new biological water-quality score system based on macroinvertebrates over a wide-range of unpolluted running-water sites. Water Res. 17, 333–347. Barbour, M.T., Yoder, C.O., 2000. The multivariate approach to bioassessment, as used in the United States of America. In: Wright, J.F., Sutcliffe, D.W., Furse, M.T. (Eds.), Assessing the Biological Quality of Fresh Waters. Freshwater Biological Association, Ambleside, Cumbria, UK, pp. 281–292. Beck, M.W., Claassen, A.H., Hundt, P.J., 2012. Environmental and livelihood impacts of dams: common lessons across development gradients that challenge sustainability. JRBM 10, 73–92. Becker, G., 1987. Net-building behaviour: tolerance and development of two caddisfly species from the river Rhine (Hydropsyche contubernalis and H. pellucidula) in relation to the oxygen content. Oecologia 73, 242–250. Bednarek, A.T., Hart, D.D., 2005. Modifying dam operations to restore rivers: ecological responses to Tennessee River dam mitigation. Ecol. Appl. 15, 997–1008. Bednarek, A.T., 2001. Undamming rivers: a review of the ecological impacts of dam removed. Environ. Manage. 27, 803–814. Belfiore, C., 1983. Efemerotteri. Consiglio Nazionale delle Ricerche, Italy. Benitez-Mora, A., Camargo, J.A., 2014. Ecological responses of aquatic macrophytes and benthic macroinvertebrates to dams in the Henares River Basin (Central Spain). Hydrobiologia 728, 167–178. Berger, E., Haase, P., Oetken, M., Sundermann, A., 2016. Field data reveal low critical chemical concentrations for river benthic invertebrates. Sci. Total Environ. 544, 864–873. Bilotta, G.S., Brazier, R.E., 2008. Understanding the influence of suspended solids on water quality and aquatic biota. Water Res. 42, 2849–2861. Blaas, H., Kroeze, C., 2016. Excessive nitrogen and phosphorus in European rivers: 2000–2050. Ecol. Indic. 67, 328–337. Bonada, N., Prat, N., Resh, V.H., Statzner, B., 2006. Devolopments in aquatic insect biomonitoring: a comparative analysis of recent approaches. Annu. Rev. Entomol. 51, 495–523. Boon, P.J., 1987. The influence of Kielder Water on trichopteran (caddisfly) populations in the River North Tyne (northern England). Regul. Rivers 1, 95–109. Brain, R.A., Hanson, M.L., Solomon, K.R., Brooks, B.W., 2008. Aquatic plants exposed to pharmaceuticals: effects and risks. Rev. Environ. Contam. Toxicol. 192, 67–115. Bruder, A., Tonolla, D., Schweizer, S.P., Vollenweider, S., Langhans, S.D., Wüest, A., 2016. A conceptual framework for hydropeaking mitigation. Sci. Total Environ. 568, 1204–1212. Bryce, S.A., Lomnicky, G.A., Kaufmann, P.R., 2010. Protecting sediment-sensitive aquatic species in mountain streams through the application of biologically based streambed sediment criteria. J. N. Am. Benthol. Soc. 29, 657–672. Céréghino, R., Cugny, P., Lavandier, P., 2002. Influence of intermittent hydropeaking on the longitudinal zonation of benthic invertebrates in a mountain stream. Int. Rev. Hydrobiol. 87, 47–60. Cañedo-Argüelles, M., Kefford, B.J., Piscart, C., Prat, N., Schäfer, R.B., Schulz, C.-J., 2013. Salinization of rivers: an urgent ecological issue. Environ. Pollut. 173, 157–167. Cairns, J., Cairn, K.L. (Eds.), 1973. Biological Methods for the Assessment of Water Quality. American Society for Testing and Materials Philadelphia. Camargo, J.A., Alonso, A., 2006. Ecological and toxicological effects of inorganic nitrogen pollution in aquatic ecosystems: a global assessment. Environ. Inter. 32, 831–849. Camargo, J.A., García de Jalón, D., 1992. Redescripción de la larva de Agapetus laniger (Pictet, 1834) (Trichoptera, Glossosomatidae) a partir de ejemplares de la Península Ibérica. Bol. Asoc. Esp. Entomol. 16, 83–93. Camargo, J.A., Voelz, N.J., 1998. Biotic and abiotic changes along the recovery gradient of two impounded rivers with different impoundment use. Environ. Monit. Assess. 50, 143–158. Camargo, J.A., Alonso, A., de la Puente, M., 2004. Multimetric assessment of nutrient enrichment in impounded rivers based on benthic macroinvertebrates. Environ. Monit. Assess. 96, 233–249. Camargo, J.A., Alonso, A., de la Puente, M., 2005. Eutrophication downstream from small reservoirs in mountain rivers of Central Spain. Water Res. 39, 3376–3384. Camargo, J.A., 1989. Estudio Ecotoxicológico del Impacto Ambiental Generado por una Regulación de Caudales y un Vertido de Flúor, sobre las Comunidades de Animales Acuáticos del Río Duratón. Doctoral Dissertation, Universidad Autónoma de Madrid,

5. Conclusions Taking account of all obtained results and their subsequent discussion, the following conclusions may be supported: (1) mitigation measures carried out in the 1990s and 2000s have markedly improved environmental conditions of the middle Duraton River, resulting in significant increases in dissolved oxygen concentrations, as well as in significant reductions of fluoride pollution and short-term flow fluctuations; (2) the community of benthic macroinvertebrates responded positively to these improvements in river environmental conditions, undergoing significant increases in abundance (total density, total biomass and EPT density) and diversity (total family richness and EPT richness) along the study area; (3) the presence of relatively sensitive benthic macroinvertebrates after mitigation measures (as indicated by increased values of BMWQ biotic indices in the 2014 sampling survey) also was the main cause for observed reductions in the environmental impact caused by disturbance points (as indicated by decreased values of the EI index), and for the observed recovering of the trophic structure of the macroinvertebrate community, with macroinvertebrate scrapers as the functional feeding group most favored; (4) structural and functional responses of benthic macroinvertebrates to mitigation measures were more marked at sampling sites that initially were more impacted (i.e., nearest to disturbance points), and less apparent at the sampling site that initially was less impacted (i.e., farthest to disturbance points); (5) improvements in river environmental conditions clearly favored the presence of Hydropsyche pellucidula and Cheumatopsyche lepida at the expense of the other hydropsychid species, particularly H. bulbifera and H. siltalai; (6) despite all monitored environmental improvements and macroinvertebrate positive responses, further mitigation and rehabilitation measures are needed to restore the ecological integrity and natural conditions of the middle Duraton River. In this regard, since the removal of Burgomillodo Dam and the cessation of the industrial activity seem improbable, additional improvements in the industrial wastewater depuration system should at least be performed, particularly to reduce high turbidity levels and sedimentation of fine inorganic matter that were negatively affecting benthic macroinvertebrates downstream from the industrial effluent. Overall, it is concluded that the multimetric approach may be used as an effective technique for assessing macroinvertebrate responses to mitigation measures in river ecosystems. 365

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los Recursos Hídricos: Sistema de Información del Anuario de Aforos. MAPAMA, Gobierno de España, Madrid. Moog, O., 1993. Quantification of daily peak hydropower effects on aquatic fauna and management to minimize environmental impacts. Regul. Rivers Res. Manage. 8, 5–14. Moretti, G., 1983. Tricotteri. Consiglio Nazionale delle Ricerche, Italy. Mueller, M., Pander, J., Geist, J., 2014. The ecological value of stream restoration measures: an evaluation on ecosystem and target species scales. Ecol. Eng. 62, 129–139. Muhar, S., Januschke, K., Kail, J., Poppe, M., Hering, D., Buijse, D.A., 2016. Evaluating good-practice cases for river restoration across Europe: context, methodological framework, selected results and recommendations. Hydrobiologia 769, 3–19. Naden, P.S., Murphy, J.F., Old, G.H., Newman, J., Scarlett, P., Harman, M., Duerdoth, C.P., Hawczak, A., Pretty, J.L., Arnold, A., Laizé, C., Hornby, D.D., Collins, A.L., Sear, D.A., Jones, J.I., 2016. Understanding the controls on deposited fine sediment in the streams of agricultural catchments. Sci. Total Environ. 547, 366–381. Nath, S.K., Dutta, R.K., 2010. Fluoride removal from water using crushed limestone. Indian J. Chem. Technol. 17, 120–125. National Water Council, 1981. River Quality: The 1980 Survey and Future Outlook. National Water Council, London. Nilsson, C., Reidy, C.A., Dynesius, M., Revenga, C., 2005. Fragmentation and flow regulation of the world’s large river systems. Science 308, 405–408. Odume, O.N., Muller, W.J., Arimoro, F.O., Palmer, C.G., 2012. The impact of water quality deterioration on macroinvertebrate communities in the Swartkops River, South Africa: a multimetric approach. Afr. J. Aquat. Sci. 37, 191–200. Palmer, M.A., Menninger, H., Bernhardt, E.S., 2010. River restoration, habitat heterogeneity and biodiversity: a failure of theory or practice? Freshw. Biol. 55, 205–222. Pan, B., Yuan, J., Zhang, X., Wang, Z., Lu, J., Yang, W., Chen, J., Li, Z., Zhao, N., Xu, M., 2016. A review of ecological restoration techniques in fluvial rivers. Int. J. Sediment Res. 31, 110–119. Petts, G.E., 1984. Impounded Rivers: Perspectives for Ecological Management. Wiley, Chichester. Räike, A., Pietiläinen, O.-P., Rekolainen, S., Kauppila, P., Pitkänen, H., Niemi, J., Raateland, A., Vuorenmaa, J., 2003. Trends of phosphorus, nitrogen and chlorophyll a concentrations in Finnish rivers and lakes in 1975–2000. Sci. Total Environ. 310, 47–59. Fundamentals of Aquatic Toxicology: Effects, Environmental Fate, and Risk Assessment. In: Rand, G.M. (Ed.), 2nd ed. Taylor and Francis, Washington, D.C. Rapport, D.J., Costanza, R., McMichael, A.J., 1998. Assessing ecosystem health. TREE 13, 397–402. Rehn, A.C., 2009. Benthic macroinvertebrates as indicators of biological condition below hydropower dams on West Slope Sierra Nevada streams California, USA. River Res. Appl. 25, 208–228. Roni, P., Beechie, T. (Eds.), 2013. Stream and Watershed Restoration: A Guide to Restoring Riverine Processes and Habitats. Wiley-Blackwell, Chichester. Rosenberg, D.M., Resh, V.H. (Eds.), 1993. Freshwater Biomonitoring and Benthic Macroinvertebrates. Chapman and Hall, London. Schindler, D.W., Vallentyne, J.R., 2008. The Algal Bowl: Overfertilization of the World's Freshwaters and Estuaries. University of Alberta Press, Edmonton, Canada. Scrimgeour, G.J., Wicklum, D., 1996. Aquatic ecosystem health and integrity: problems and potential solutions. J. N. Am. Benthol. Soc. 15, 254–261. Seidel, M., Lüderitz, V., 2015. Monitoring of river restoration success: comparison between a multimetric approach and the evaluation according to the Water Framework Directive. Wasserwirtschaft 105, 24–28. Serrano Balderas, E.C., Grac, C., Berti-Equille, L., Armienta Hernandez, M.A., 2016. Potential application of macroinvertebrates indices in bioassessment of Mexican streams. Ecol. Indic. 61, 558–567. Smith, V.H., Tilman, G.D., Nekola, J.C., 1999. Eutrophication: impacts of excess nutrient inputs on freshwater, marine, and terrestrial ecosystems. Environ. Pollut. 100, 179–196. Sokal, R.R., Rohlf, F.J., 1995. Biometry: The Principles and Practice of Statistics in Biological Research, 3rd ed. Freeman, New York. Stehle, S., Schulz, R., 2015. Agricultural insecticides threaten surface waters at the global scale. Proc. Natl. Acad. Sci. U. S. A. 112, 5750–5755. Sundermann, A., Antons, C., Cron, N., Lorenz, A.W., Hering, D., Haase, P., 2011. Hydromorphological restoration of running waters: effects on benthic invertebrate assemblages. Freshw. Biol. 56, 1689–1702. Suter, G.W., 1993. A critique of ecosystem health concepts and indexes. Environ. Toxicol. Chem. 12, 1533–1539. Tachet, H., Bournaud, M., Richoux, P., 1981. Introduction à l’Etude des Macroinvertebrés des Eaux Douces. L’Association Française de Limnologie, Paris. Tachet, H., Richoux, P., Bournaud, M., Usseglio-Polatera, P., 2003. Invertebrés d’Eau Douce (Systématique, Biologie, Écologie). CNRS Editions, Paris. Thorp, J.H., Covich, A.P. (Eds.), 2010. Ecology and Classification of North American Freshwater Invertebrates, 3rd ed. Academic Press, San Diego. Turley, M.D., Bilotta, G.S., Chadd, R.P., Extence, C.A., Brazier, R.E., Burnside, N.G., Pickwell, A.G.G., 2016. A sediment-specific family-level biomonitoring tool to identify the impacts of fine sediment in temperate rivers and streams. Ecol. Indic. 70, 151–165. Verberk, W.C.E.P., Leuven, R.S.E.W., van Duinen, G.A., Esselink, H., 2010. Loss of environmental heterogeneity and aquatic macroinvertebrate diversity following largescale restoration management. Basic Appl. Ecol. 11, 440–449. Verdonschot, R.C.M., Kail, J., McKie, B.G., Verdonschot, P.F.M., 2016. The role of benthic microhabitats in determining the effects of hydromorphological river restoration on macroinvertebrates. Hydrobiologia 769, 55–66. Waghmare, S.S., Arfin, T., 2015. Fluoride removal from water by calcium materials: a state-of-the-art review. IJIRSET 4, 8090–8102.

España. Camargo, J.A., 1990. Performance of a new ecotoxicological index to assess environmental impacts on freshwater communities. Bull. Environ. Contam. Toxicol. 48, 428–434. Camargo, J.A., 1993. Macrobenthic surveys as a valuable tool for assessing freshwater quality in the Iberian Peninsula. Environ. Monit. Assess. 24, 71–90. Camargo, J.A., 1994. The importance of biological monitoring for the ecological risk assessment of freshwater pollution: a case study. Environ. Int. 20, 229–238. Camargo, J.A., 2003. Fluoride toxicity to aquatic organisms: a review. Chemosphere 50, 251–264. Casper, A.F., Thorp, J.H., Davies, S.P., Courtemanch, D.L., 2006. Ecological responses of zoobenthos to dam removal on the Kennebec River Maine, USA. Arch. Hydrobiol. Suppl. 158, 541–555. Chapman, J.M., Proulx, C.I., Veilleux, M.A.N., Levert, C., Bliss, S., André, M.-E., Lapointe, N.W.R., Cooke, S.J., 2014. Clear as mud: a meta-analysis on the effects of sedimentation on freshwater fish and the effectiveness of control measures. Water Res. 56, 190–202. Consiglio, C., 1980. Plecotteri. Consiglio Nazionale delle Ricerche, Italy. Darby, S., Sear, D. (Eds.), 2008. River Restoration: Managing the Uncertainty in Restoring Physical Habitat. Wiley, Chichester. Debenest, T., Silvestre, J., Coste, M., Pinelli, E., 2010. Effects of pesticides on freshwater diatoms. Rev. Environ. Contam. Toxicol. 203, 87–103. Della Rossa, P., Jannoyer, M., Mottes, C., Plet, J., Bazizi, A., Arnaud, L., Jestin, A., Woignier, T., Gaude, J.M., Cattan, P., 2017. Linking current river pollution to historical pesticide use: insights for territorial management? Sci. Total Environ. 574, 1232–1242. Edington, J.M., Hildrew, A.G., 1981. A Key to the Caseless Caddis Larvae of the British Isles, with Notes on Their Ecology. Freshwater Biological Association Scientific Publication no 43. Engels, S., Neumann, D., Loebbel, H., Bruehne, M., 1996. Waiting for Hydropsyche: why has only one of at least four local Hydropsyche species returned into the Lower River Rhine? Arch. Hydrobiol. Suppl. 113, 313–317. Englund, G., Malmqvist, B., Zhang, Y.X., 1997. Using predictive models to estimate effects of flow regulation on net-spinning caddis larvae in North Swedish rivers. Freshw. Biol. 37, 687–697. Ernst, A.G., Warren, D.R., Baldigo, B.P., 2012. Natural-channel-design restorations that changed geomorphology have little effect on macroinvertebrate communities in headwater streams. Restor. Ecol. 20, 532–540. Extence, C.A., Bates, A.J., Forbes, W.J., Barham, P.J., 1987. Biologically based water quality management. Environ. Pollut. 45, 221–236. Fernández-Lop, A., 1987. Análisis de la Variabilidad Intra e Interespecífica en el Grupo Echinogammarus berilloni (Crustacea, Amphipoda). Doctoral Dissertation, Universidad Complutense de Madrid, España. Fore, L.S., Karr, J.R., Wisseman, R.W., 1996. Assessing invertebrate responses to human activities: evaluating alternative approaches. J. N. Am. Benthol. Soc. 15, 212–231. Green, J., Trett, M.W. (Eds.), 1989. The Fate and Effects of Oil in Freshwater. Elsevier Applied Science, London. Harper, D.M., 1992. Eutrophication of Freshwaters. Principles, Problems and Restoration. Chapman and Hall, London. Hauer, F.R., Lamberti, G.A. (Eds.), 1996. Methods in Stream Ecology. Academic Press, San Diego. Hauer, C., Holzapfel, P., Leitner, P., Graf, W., 2017. Longitudinal assessment of hydropeaking impacts on various scales for an improved process understanding and the design of mitigation measures. Sci. Total Environ. 575, 1503–11514. Hellawell, J.M., 1986. Biological Indicators of Freshwater Pollution and Environmental Management. Elsevier Applied Science Publisher, London. Henley, W.F., Patterson, M.A., Neves, R.J., Lemly, A.D., 2000. Effects of sedimentation and turbidity on lotic food webs: a concise review for natural resource managers. Rev. Fish Sci. 8, 125–139. Jones, J.I., Murphy, J.F., Collins, A.L., Sear, D.A., Naden, P.S., Armitage, P.D., 2012. The impact of fine sediment on macro-invertebrates. River Res. Appl. 28, 1055–1071. Kail, J., Brabec, K., Poppe, M., Januschke, K., 2015. The effect of river restoration on fish, macroinvertebrates and aquatic macrophytes: a meta-analysis. Ecol. Indic. 58, 311–321. Karr, J.R., 1991. Biological integrity: a long-neglected aspect of water resource management. Ecol. Appl. 1, 66–84. Klemm, D.J., Blocksom, K.A., Thoeny, W.T., Fulk, F.A., Herlihy, A.T., Kaufmann, P.R., Cormier, S.M., 2002. Methods development and use of macroinvertebrates as indicators of ecological conditions for streams in the Mid-Atlantic Highlands Region. Environ. Monit. Assess. 78, 169–212. Lüderitz, V., Speierl, T., Langheinrich, U., Völkl, W., Gersberg, R.M., 2011. Restoration of the upper Main and Rodach rivers: the success and its measurement. Ecol. Eng. 37, 2044–2055. Lauters, F., Lavandier, P., Lim, P., Sabaton, C., Belaud, A., 1996. Influence of hydropeaking on invertebrates and their relationship with fish feeding habits in a Pyrenean river. Regul. Rivers Res. Manage. 12, 563–573. Lessard, J.L., Hayes, D.B., 2003. Effects of elevated water temperature on fish and macroinvertebrate communities below small dams. River Res. Appl. 19, 721–732. Lorenz, A.W., Korte, T., Sundermann, A., Januschke, K., Haase, P., 2012. Macrophytes respond to reach-scale river restorations. J. Appl. Ecol. 49, 202–212. Louhi, P., Mykra, H., Paavola, R., Huusko, A., Vehanen, T., Maki-Petays, A., Muotka, T., 2011. Twenty years of stream restoration in Finland: little response by benthic macroinvertebrate communities. Ecol. Appl. 21, 1950–1961. Merritt, R.W., Cummins, K.W., Berg, M.B. (Eds.), 2008. An Introduction to the Aquatic Insects of North America, 4th ed. Kendall/Hunt Publishing Company, Dubuque, Iowa. Ministerio de Agricultura y Pesca, Alimentación y Medio Ambiente, 2017. Evaluación de

366

Ecological Indicators 83 (2017) 356–367

J.A. Camargo

Wood, P.J., Armitage, P.D., 1997. Biological effects of fine sediment in the lotic environment. Environ. Manage. 21, 203–217. World Health Organization, 2002. Environmental Health Criteria 227: Fluorides. WHO, Geneva, Switzerland. Zamora-Muñoz, C., Alba-Tercedor, J., García de Jalón, D., 1995. The larvae of the genus Hydropsyche (Hydropsychidae; Trichoptera) and key for the identification of species of the Iberian Peninsula. Mitt. Schweiz. Entomol. Ges. 68, 189–210. Ziglio, G., Siligardi, M., Flaim, G. (Eds.), 2006. Biological Monitoring of Rivers. Wiley, Chichester.

Walling, D.E., Fang, D., 2003. Recent trends in the suspended sediment loads of the world’s rivers. Global Planet. Change 39, 111–126. Walther, D.A., Whiles, M.R., 2008. Macroinvertebrate responses to constructed riffles in the Cache River, Illinois, USA. Environ. Manage. 41, 516–527. Ward, J.V., Stanford, J.A. (Eds.), 1979. The Ecology of Regulated Streams. Plenum Press, New York. Washington, H.G., 1984. Diversity, biotic and similarity indices: a review with special relevance to aquatic ecosystems. Water Res. 18, 653–694. Wetzel, R.G., Likens, G.E., 2000. Limnological Analyses, 3rd ed. Springer, New York.

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