Hydrobiologia (2014) 728:167–178 DOI 10.1007/s10750-014-1816-6
PRIMARY RESEARCH PAPER
Ecological responses of aquatic macrophytes and benthic macroinvertebrates to dams in the Henares River Basin (Central Spain) Alfonso Benı´tez-Mora • Julio A. Camargo
Received: 21 August 2013 / Revised: 23 December 2013 / Accepted: 27 January 2014 / Published online: 8 February 2014 Ó Springer International Publishing Switzerland 2014
Abstract The ecological responses of aquatic macrophytes and benthic macroinvertebrates to deeprelease dams in three impounded rivers of the Henares River Basin (Central Spain) were studied, specially focusing on the effects of nutrient enrichment caused by deep releases on these two freshwater communities. Three sampling sites, one upstream and two downstream from the reservoir, were established in each impounded river. Sampling surveys to collect submersed macrophytes and benthic macroinvertebrates at each sampling site were carried out in spring– summer of 2009 and 2011. Water temperature tended to decrease downstream from dams, whereas nitrate and phosphate concentrations tended to increase. These abiotic changes, particularly the downstream nutrient enrichment, apparently affected the macrophyte and macroinvertebrate communities. In the case of submersed macrophytes, total coverage and taxa richness increased downstream from dams. In the case of benthic macroinvertebrates, total density and total biomass also increased downstream, but taxa richness tended to decrease. Scrapers appeared to be the macroinvertebrate feeding group most favored
Handling editor: Sidinei Magela Thomaz A. Benı´tez-Mora J. A. Camargo (&) Unidad Docente de Ecologı´a, Departamento de Ciencias de la Vida, Universidad de Alcala´, 28805 Alcala´ de Henares, Madrid, Spain e-mail:
[email protected]
downstream from dams as a probable consequence of the positive effect of nutrient enrichment on periphyton and perilithon abundance. Nutrients would ultimately come from water runoff over agricultural lands and over semi-natural forests and pastures, being subsequently accumulated in the hypolimnion of reservoirs. Keywords Submersed macrophytes Benthic macroinvertebrates Ecological responses Deep-release dams Central Spain
Introduction Since the earliest civilizations were established on the Indus, Nile, and Tigris-Euphrates Rivers, man has tried to control and regulate the flow of rivers for the benefit of human interests such as water supply, irrigation, effective flood control, navigation, inexpensive and efficient power generation, and recreational opportunities. 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 80% of the total discharge of large rivers in the northern third of the world being impacted by river regulation (Dynesius & Nilsson, 1994; McCully, 1996; Bednarek, 2001). Unfortunately, dam construction and operation and reservoir development can cause important adverse effects on the structure of freshwater communities by modifying
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physicochemical conditions upstream, downstream, and in reservoirs (Ward & Stanford, 1979; Petts, 1984; Graf, 1999; World Commission on Dams, 2000; Nilsson et al., 2005; Beck et al., 2012). In addition, reservoirs can contribute to the atmospheric greenhouse effect through the decomposition of organic matter in their sediments and the subsequent emission to the atmosphere of greenhouse gases, mostly carbon dioxide and methane (St. Louis et al., 2000; Ma¨kinen & Khan, 2010; Barros et al., 2011). Although it is evident that specific effects of dams on the physical, chemical, and biological components of fluvial ecosystems depend greatly on the size and type of dams, dam location along the river, and ecological characteristics of each impounded reach, numerous field studies around the world have extensively documented and identified several general impacts. Reductions in river flow and matter transport are a direct consequence of transforming the natural lotic ecosystem (the river) into an artificial lentic habitat (the reservoir), where suspended particles tend to settle and inorganic nutrients (like phosphorous) may be trapped and accumulated in the hypolimnion and sediments (but with potential nutrient resolubilization) (Petts, 1984; Kondolf, 1997; World Commission on Dams, 2000; Poff & Zimmerman, 2010; Beck et al., 2012). The entrapment and accumulation of nutrients can promote eutrophication within reservoirs, but it can lead to oligotrophication downstream, particularly if dams release surface waters, causing reductions in fisheries resources (Petts, 1984; Ney, 1996; Stockner et al., 2000). Dams also fragment the continuity of rivers and hamper the fluvial migration of fish and other aquatic organisms, with freshwater biota being able to be harmed by turbines of hydropower impoundments as it attempts to pass (Drinkwater & Frank, 1994; Nilsson et al., 2005; Meixler et al., 2009; Dudgeon, 2010). On the other hand, significant changes in water temperature and dissolved oxygen, both within reservoirs and downstream from dams, may cause important adverse effects on the abundance and diversity of zoobenthic communities (Ward & Stanford, 1979; Petts, 1984; Armitage et al., 1987; Camargo & Voelz, 1998; Lessard & Hayes, 2003; Rehn, 2009). Further environmental problems can arise when hydroelectric power generation induces short-term flow fluctuations downstream, causing negative impacts on aquatic organisms that depend on critical thresholds of water level, velocity, or timing for life history stages (Ward & Stanford, 1979;
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Petts, 1984; Armitage et al., 1987; Camargo & Voelz, 1998; Poff & Zimmerman, 2010; Beck et al., 2012). Little attention has however been paid to the fact that impounded reaches of rivers and streams can experience eutrophication downstream as a consequence of deep releases from dams (Camargo et al., 2005). Ward & Stanford (1983) suggested that, if the water residence time in impounded reaches is increased by the presence of dams, deep-release impoundments could conceivably raise nutrient levels downstream and, consequently, cause eutrophication. More recently, Casas et al. (2000) and Camargo et al. (2005) reported increased nutrient concentrations below dams in the impounded upper reaches of five Spanish rivers. Furthermore, Camargo et al. (2005) found that periphyton chlorophyll a and macroinvertebrate scrapers increased significantly downstream as a likely consequence of nutrient enrichment (phosphate, primarily) caused by deep releases from dams. Other studies have also showed that the damming of upper reaches can impact the activity of macroinvertebrate shredders in relation to leaf litter breakdown (Mene´ndez et al., 2012; Gonza´lez et al., 2013). In this research, we examine the ecological responses of aquatic macrophytes and benthic macroinvertebrates to deep-release dams in the impounded middle reaches of three rivers of the Henares River Basin (Central Spain). We focus specially on the effects of the potential nutrient enrichment caused by deep releases from dams on these two freshwater communities, taking into account the two following major hypotheses: 1) because nutrient enrichment often affects positively the development of primary producers, macrophyte coverage should increase downstream from dams as a direct response to nutrient enrichment; 2) macroinvertebrate abundance, particularly the abundance of macroinvertebrate scrapers, should increase downstream from dams as a response to the positive effect of nutrient enrichment on primary producers.
Materials and methods The study impounded reaches and sampling sites Field studies were conducted in the impounded middle reaches of three rivers (Bornova, Can˜amares, and Sorbe Rivers) of the Henares River Basin (Central Spain), within the Tajo River Basin (Fig. 1). The
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169 Table 1 Characteristics of reservoirs/dams in the study impounded rivers River Reservoir/dam
Bornova Alcorlo
Can˜amares Pa´lmaces
Sorbe Belen˜a
Year construction
1978
1954
1982
Reservoir area (km2) Dam height (m)
5.99
2.70
2.45
73
43
57
Reservoir capacity (hm3)
180
31
53
Trophic status
Oligotrophic– mesotrophic
Mesotrophic
Oligotrophic
Reservoir use
Water supply, irrigation
Irrigation
Water supply
Data from Confederacio´n Hidrogra´fica del Tajo (2012)
Fig. 1 The study impounded rivers (Sorbe, Bornova and Can˜amares Rivers) in the Henares River Basin (Central Spain), within the Tajo River Basin. Three sampling sites (one upstream and two downstream from the dams) were established in each impounded river. Land use was established according to Camargo (2006) and Go´mez-Sal (2011), following the CORINE Agricultural Land Cover. White empty areas indicate seminatural forests
natural flow regime of these rivers is characterized by maximum flows during the winter and spring seasons and minimum flows during the summer and autumn seasons (Camargo, 2006). A deep-release dam (43–73 m in height; Table 1) is found in the middle reaches of each impounded river, forming three storage reservoirs (Fig. 1; Table 1): Belen˜a Reservoir in Sorbe River, Pa´lmaces Reservoir in Can˜amares River, and Alcorlo Reservoir in Bornova River. According to Confederacio´n Hidrogra´fica del Tajo (2012), these reservoirs may be classified as oligotrophic and mesotrophic, being basically used for water supply and irrigation (Table 1).
Three sampling sites (river depth \80 cm), one upstream and two downstream from the reservoir, were established in each impounded river (Fig. 1): S1, S2, and S3 in Sorbe River; B1, B2, and B3 in Bornova River; C1, C2, and C3 in Can˜amares River. Upstream sampling sites (B1, C1, S1) were situated between 0.8 and 1.5 km upstream from dams, first downstream sampling sites (B2, C2, S2) were situated between 0.4 and 0.7 km downstream from dams, and second downstream sampling sites (B3, C3, S3) were situated between 1.5 and 2.2 km downstream from dams. The river substrate was mainly stony, with gravels and boulders at all sampling sites. According to Camargo (2006) and Go´mez-Sal (2011), the land around the study areas is mainly dedicated to rain-fed agriculture and semi-natural forests (Fig. 1), with some irrigation land being significant along the river valleys of Bornova and Can˜amares Rivers. In addition, upstream and downstream from reservoirs, the riparian forest is relatively narrow and mainly composed by willows (Salix alba, S. eleagnos and S. purpurea), poplars (Populus nigra and P. alba), and elms (Ulmus minor) (Camargo, 2006; Go´mez-Sal, 2011). On the other hand, the major anthropogenic point sources causing freshwater eutrophication (e.g., wastewaters from intensive animal farming, industrial and municipal effluents) were lacking, and no tributary stream occurred between upstream and downstream sampling sites (Camargo, 2006; Go´mez-Sal, 2011). All this information was also
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verified from direct visual examination during our sampling surveys. Water sampling and analyses Since its foundation in 1926, the Confederacio´n Hidrogra´fica del Tajo (CHT) is the official institution in charge of the integrated basin-based water management of the Tajo River Basin, and its Water Quality Control Network has been progressively developed from the early 1960s to the present. The Water Quality Control Network of CHT provided us with a raw data set of the following physicochemical variables recorded at each sampling site in 2009 and 2010 (Confederacio´n Hidrogra´fica del Tajo, 2012): water temperature, dissolved oxygen, pH, nitrate, and phosphate. These physiochemical parameters were measured following standard methods (American Public Health Association, 1998). In addition, we measured in situ river current/water velocity with a helical gear current meter in accordance with Wetzel & Likens’ (2000) indications. Macrophyte and macroinvertebrate sampling and analyses Two sampling surveys to examine the communities of submersed macrophytes and benthic macroinvertebrates were carried out in May, June, and July of 2009 and 2011. To sample submersed macrophytes, an area of about 100 m2 was selected at each sampling site, and coverage percentage was estimated for each taxon and for the whole community. Briefly, three river transects were conducted at each sampling site, on each sampling survey, examining five sampling squares (45 9 45 cm subdivided in four equal parts) per transect (Hauer & Lamberti, 1996). The five sampling squares were randomly chosen subsamples of each river transect. The identification to the genus or species level followed Cirujano & Medina (2002) and Casas et al. (2006). To sample benthic macroinvertebrates, a hand net with a 250 lm mesh size (Hauer & Lamberti, 1996) was used to collect four riffle bottom samples at each sampling site on each sampling survey. For each bottom sample, an area of about 1 m2 was sampled, including macroinvertebrates associated with submersed macrophytes. All macroinvertebrate samples were preserved in 4% formalin until laboratory analyses. In the laboratory, benthic macroinvertebrates were identified
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and counted with a light stereomicroscope. Taxonomic identification to the family level followed Tachet et al. (2003). However, in some cases, we identified macroinvertebrates to the genus or species levels. After identification and counting, macroinvertebrate samples were dried in an oven at 60°C for 48 h, and weighed with a precision scale to obtain macroinvertebrate biomass (dry weight) at each sampling site. To examine changes in the trophic structure of the macrobenthic community, benthic macroinvertebrates were allocated to five functional feeding groups in accordance with Tachet et al. (2003) and Thorp & Covich (2010): shredders basically feed on coarse particulate organic matter; scrapers mainly feed on periphyton and perilithon; collector-gatherers feed on fine organic detritus, but many of them can also feed on periphyton and perilithon; collector-filterers mainly feed on organic material suspended in the water column; and predators feed on animal preys. The percentage contribution of each functional feeding group at each sampling site was calculated on the basis of biomass estimates. Biological metrics Abundance and diversity metrics were estimated for the whole macrophyte and macroinvertebrate communities: macrophyte abundance is expressed as the total coverage percentage per river transect; macrophyte diversity or richness is expressed as the total number of taxa (genera or species) per river transect; macroinvertebrate abundance is expressed as the total density (number of individuals) per square meter of benthic substratum, and also as the total biomass (dry weight) per square meter of benthic substratum; macroinvertebrate diversity or richness is expressed as the total number of families per sampling unit (the surber sampler). Significant changes in the value of these metrics can indicate environmental disturbances, including freshwater pollution and nutrient enrichment (Washington, 1984; Hellawell, 1986; Rosenberg & Resh, 1993; Schneider & Melzer, 2003; Sua´rez et al., 2005; Szoszkiewicz et al., 2006; Camargo et al., 2011). Statistical analyses All statistical analyses were performed using the STATISTICA 7.0 package. Mean values of physicochemical parameters were compared between habitat sampling sites (B1, B2, B3; C1, C2, C3; S1, S2, S3) by
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means of one way analysis of variance (ANOVA). When significant differences between habitat sampling sites were found, each downstream sampling site was compared with the respective upstream sampling site (B1, C1 or S1) by means of a parametric post hoc Dunnett test (Sokal & Rohlf, 1995). The same statistical approach was performed for mean values of biological metrics, including mean abundances of macrophytes and macroinvertebrates. In order to reduce the magnitude of the type-I error, as several univariate comparisons were conducted, a significance level of P \ 0.01 was selected (Sokal & Rohlf, 1995). On the other hand, ordination analysis was used to identify environmental gradients and their relation to taxonomic composition and metrics. A detrended correspondence analysis (DCA) was carried out to decide whether it is appropriate to use linear or unimodal ordination methods. This analysis revealed a gradient length of the first axis of 1.582 SD-units for macroinvertebrates and 3.916 SD-units for macrophytes. This suggests that the use of lineal ordination methods is appropriate for this data analysis (Lepsˇ & Sˇmilauer, 2003). We subsequently carried out principal components analysis (PCA) on biotic data (macroinvertebrate and macrophytes) and water physicochemical data to identify the main ecological gradients. The analyses were performed separately for macroinvertebrate and macrophytes. DCA and PCA were realized using CANOCO version 4.5.
Results Water velocity, pH, and dissolved oxygen did not exhibit clear trends along the study impounded reaches (Table 2). In contrast, water temperature tended to decrease downstream from dams, whereas nutrient concentrations tended to increase (Table 2): phosphate concentrations were higher at downstream sampling sites 1 (B2, C2, S2), and nitrate concentrations were higher at downstream sampling sites 2 (B3, C3, S3). Macrophyte total coverage (except in Can˜amares River) and macrophyte taxa richness were significantly (P \ 0.01) higher at downstream sampling sites than at upstream sampling sites (Table 3). Apium nodiflorum, Groenlandia densa, and Potamogeton pectinatus were never found at B1, C1, and S1 (Table 4). G. densa coverages were dominant at downstream sampling sites C2 and S2, whereas P.
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pectinatus coverages were dominant at downstream sampling sites B2, C3, and S3. The aquatic moss Leptodictyum riparium was only found at Bornova River (Table 4) and its coverage values were similar at upstream (B1) and downstream (B2) sampling sites, but significantly higher (P \ 0.01) at downstream sampling site B3. The aquatic moss Fontinalis antipyretica was only found in Can˜amares River (Table 4) and its coverage values decreased at downstream sampling sites (C2 and C3). Lastly, the green macroalga Cladophora sp. was found in the three study areas (Table 4), its coverage increasing downstream from dams in Sorbe and Bornova Rivers. However, in Can˜amares River, Cladophora coverage was significantly higher at the reference site (C1) than at downstream sampling sites (C2 and C3). Total density and total biomass of the macroinvertebrate community were significantly higher at downstream sampling sites of the three impounded reaches (Table 3), whereas macroinvertebrate taxa richness (number of families) tended to decrease (except at S3) (Table 3). In general, stoneflies (Leuctridae, Perlidae, and Taeniopterygidae families) and caddisflies (Hydropsychidae and Limnephilidae families) were the macroinvertebrates most adversely affected, their abundances tending to decrease downstream from dams (Table 5). Conversely, aquatic snails (Ancylus fluviatilis, particularly), amphipods (Echinogammarus echinosetosus), midge larvae (Chironomidae and Simuliidae families), aquatic beetles (Elmidae family), planarians (Planariidae family), and some ephemeropterans (Baetis rhodani, particularly) were the macroinvertebrates most favored, their abundances tending to increase at downstream sampling sites (Table 5). At upstream sampling sites, macroinvertebrate filter feeders were dominant in Bornova (B1) and Sorbe (S1) Rivers (Fig. 2) because of the relatively high biomass (dry weight) of Hydropsychidae and Simuliidae larvae, whereas macroinvertebrate shredders were dominant in Can˜amares (C1) River (Fig. 2) as a consequence of the relatively high biomass of stoneflies, caddisflies, and amphipods. However, at downstream sampling sites B2, C2, and S2, macroinvertebrate scrapers were the dominant feeding group (Fig. 2) because of the increased abundances of ephemeropterans, aquatic snails, and beetles. At downstream sampling sites B3, C3, and S3, shredders recovered, being the dominant feeding group (Fig. 2) as a consequence of the increased abundance of gammarids. Filter-feeders also
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Raw data were obtained from Confederacio´n Hidrogra´fica del Tajo (2012)
* Significant differences (P \ 0.01) at downstream sites (B2, B3; C2, C3; S2, S3) with regard to the upstream sites (B1, C1, S1) after ANOVA and a post hoc Dunnett test
74.5 ± 19.0 121.5 ± 56.6* 53.8 ± 4.8 67.5 ± 7.4 111.0 ± 61.6 58.3 ± 2.5 92.0 ± 42.9 Phosphate (lg/l)
60.8 ± 8.3
78.0 ± 22.0
2.8 ± 0.5* 1.7 ± 0.8 0.9 ± 0.1 7.0 ± 4.2* 2.6 ± 1.0 1.9 ± 1.5 2.4 ± 1.9 1.0 ± 0.1 Nitrate (mg/l)
7.9 ± 0.3
8.8 ± 6.7*
9.9 ± 1.2
7.6 ± 0.2 7.4 ± 0.2
10.2 ± 0.6 8.3 ± 1.1
7.0 ± 0.4 7.9 ± 0.2
10.5 ± 0.6 10.3 ± 1.5
8.2 ± 0.1 8.0 ± 0.1
10.0 ± 1.5 10.9 ± 1.3
8.1 ± 0.3 pH
10.8 ± 1.0
7.7 ± 0.1
12.3 ± 3.9
Dissolved oxygen (mg/l)
10.6 ± 0.9
17.0 ± 8.0*
13.1 ± 3.4 16.8 ± 2.0
62.5 ± 23.0 59.3 ± 27.5
13.0 ± 3.7 12.5 ± 7.0
44.5 ± 14.6 31.1 ± 11.9
14.9 ± 4.6 11.4 ± 4.5
66.5 ± 31.0
12.2 ± 5.3
54.0 ± 18.4
14.9 ± 3.3
63.9 ± 13.9
Water temperature (°C)
Water velocity (cm/s)
S2 S1 C1 B2
B3 B1
C2
C3
Sorbe River Can˜amares River Bornova River
Table 2 Mean (n = 4, n = 20 only for water velocity; ±SD) values of water physicochemical parameters at each sampling site for 2009 and 2010 years
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35.3 ± 16.2
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S3
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recovered in part at B3 and S3 (Fig. 2) as a consequence of the increased abundance of Simuliidae larvae. Principal component analyses with submersed macrophytes and water physicochemical parameters showed that the first two axes of the PCA accounted for 72.5% of the total variance, with axes 1 and 2 explaining 46.6 and 25.9% of the total variance, respectively (Fig. 3). The first axis was positively associated with water temperature, and negatively associated with water velocity and nitrates (Fig. 3). On this axis, F. antipyretica and Cladophora sp. occupied a positive position, whereas A. nodiflorum, P. pectinatus, and L. riparium occupied a negative position (Fig. 3). The second axis was highly positively correlated with phosphates and negatively associated with water velocity (Fig. 3). The phanerogams G. densa, M. spicatum, and R. peltatus were positively associated with the second axis, while the aquatic moss L. riparium was negatively associated (Fig. 3). Principal component analyses with macroinvertebrate feeding groups and water physicochemical parameters showed that the first two axes of the PCA accounted for 93% of the total variance, with axes 1 and 2 explaining 63 and 30% of the total variance, respectively (Fig. 4). The first axis was clearly dominant, being positively associated with phosphates, and negatively associated with water velocity and water temperature (Fig. 4). Upstream sampling sites (B1, C1, S1) and downstream sampling sites 1 (B2, C2, S2) were clearly associated along the first axis, with downstream sites occupying a positive position and upstream sites on the opposite side. Macroinvertebrate scrapers and collector-gatherers were positively associated with the first axis, whereas macroinvertebrate predators and filter-feeders were negatively associated. In contrast, the second axis was slightly correlated with phosphates and water temperature, and did not discriminate clearly between sampling sites (Fig. 4). Macroinvertebrate shredders were clearly associated with downstream sampling sites 2 (B3, C3, S3).
Discussion Alterations in water velocity and flow, and significant decreases in water temperature and dissolved oxygen, have been previously observed downstream from deep-release dams in impounded rivers and streams around the world (Ward & Stanford, 1979; Petts,
* Significant differences (P \ 0.01) at downstream sites (B2, B3; C2, C3; S2, S3) with regard to the upstream sites (B1, C1, S1) after ANOVA and a post hoc Dunnett test
3 ± 0*
57.5 ± 4.9* 64.7 ± 4.1*
2 ± 0* 1±0
4.7 ± 1.6 58.3 ± 5.7
6 ± 0.6* 6 ± 0.4*
63.5 ± 7.0 68.0 ± 1.6
2±0 3 ± 0*
48.8 ± 14.0* 47.5 ± 5.6*
5 ± 0* 1±0
5.3 ± 1.4 Macrophyte total coverage
Macrophyte richness (gen. or sp.)
2.6 ± 0.4* 3.0 ± 0.3* 1.1 ± 0.1 3.5 ± 0.3* 3.7 ± 0.1* 2.1 ± 0.1 3.4 ± 0.4* 2.5 ± 0.2
4.7 ± 0.4*
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Macroinvertebrate biomass (g/m2)
21.0 ± 0.7* 16.0 ± 0.4 17.0 ± 1.1 14.0 ± 1.0* 17.0 ± 1.1*
18.0 ± 0.7
16.0 ± 0.8 19.0 ± 1.0 Macroinvertebrate richness (fam.)
13.0 ± 1.0*
1,180 ± 155* 2,254 ± 258* 808 ± 93.4 1,592 ± 204* 1,754 ± 116*
474 ± 28.9 1,387 ± 86.6* 901 ± 75.4 Macroinvertebrate density (individ./m2)
1,484 ± 77.4*
S3 S2 S1 C1 B3 B2 B1
C2
C3
Sorbe River Can˜amares River Bornova River
Table 3 Mean (n = 6–8; ±SD) values of biological metrics for the macrobenthic and macrophytic communities at each sampling site for 2009 and 2011 sampling surveys
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1984; Voelz & Ward, 1991; Camargo & Voelz, 1998; Ogbeibu & Oribhabor, 2002; Rehn, 2009; Poff & Zimmerman, 2010). Furthermore, although the entrapment and accumulation of nutrients within reservoirs can lead to oligotrophication downstream, particularly if dams release surface waters (Petts, 1984; Ney, 1996; Stockner et al., 2000), some field studies have reported increased nutrient concentrations below deep-release dams (Casas et al., 2000; Camargo et al., 2005). In the present research, mean values of water velocity, water temperature, and dissolved oxygen were not significantly (P [ 0.01) different at downstream sampling sites with regard to upstream sampling sites (Table 2). It however was evident that water temperature tended to decrease downstream from dams at all study impounded reaches (Table 2). On the other hand, the significant (P \ 0.01) increases in inorganic nutrients (nitrate and phosphate) at downstream sampling sites (Table 2) give support to the hypothesis that impounded middle reaches can experience nutrient enrichment as a consequence of deep releases from dams. Because many agricultural lands are found in the study areas, in addition to seminatural forests and pastures which are often used to sustain livestock (cows and horses), we consider that nutrients and organic matter would ultimately come from land/forest runoff, being subsequently accumulated in the hypolimnion of reservoirs. Numerous studies have previously shown that water runoff over agricultural lands and over forests and pastures can significantly increase the amount of nutrients and organic matter in recipient rivers (Mesters, 1995; Wetzel, 2001; Hilton et al., 2006; Torrent et al., 2007; Yang et al., 2010; Steffen et al., 2013). Nevertheless, we consider that, if terrestrial vegetation was not completely removed before filling reservoirs, this fact could not contribute currently to the observed downstream nutrient enrichment, owing to the long time passed since dams were constructed and reservoirs developed (Table 1). The observed downstream nutrient enrichment could in part be responsible for the significant increases in macrophyte richness and coverage downstream from dams (Table 3). Indeed, the downstream increases in the coverage of Groenlandia densa, Ranunculus peltatus, Potamogeton pectinatus and Apium nodiflorum would denote the nutrient enrichment caused by deep releases from dams (Table 4; Fig. 3). Increases in the abundance
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Table 4 Mean (n = 6) relative abundances of submersed macrophytes at sampling sites of each impounded river Can˜amares River
Bornova River B1
B2
B3
Sorbe River
C1
C2
C3
S1
S2
S3
Apium nodiflorum
–
15.4
13.7
–
11.3
15.7
–
Cladophora sp.
–
10.2
–
83.8
15.2
11.1
100
–
–
39.7
20.6
Fontinalis antipyretica
–
–
–
16.2
11.5
7.1
–
–
–
Groenlandia densa
–
22.5
–
–
33.1
19.7
–
60.3
28.7
Leptodictyum riparium
100
15.4
56.0
–
–
–
–
–
–
Myriophyllum spicatum
–
–
–
–
12.9
–
–
–
–
Potamogeton pectinatus
–
36.5
30.4
–
–
25.1
–
–
50.7
Ranunculus peltatus
–
–
–
–
16.0
21.1
–
–
–
Relative abundances were calculated on the basis of coverage estimates (%)
and diversity of primary producers, including submersed macrophytes, have been observed in other rivers and streams experiencing nutrient enrichment (Cirujano & Medina, 2002; Schneider & Melzer, 2003; Sua´rez et al., 2005; Szoszkiewicz et al., 2006; Camargo et al., 2011). Nevertheless, in other cases, the nutrient enrichment of freshwater ecosystems by nitrogen and phosphorus compounds has caused profound shifts in the macrophyte communities via the replacement of rooted macrophytes by green macroalgae or phytoplankton (Marques et al., 2003; Camargo et al., 2005; Steffen et al., 2013). In our study impounded reaches, the coverage of Cladophora sp. seemed to be highly sensitive to water temperature and water velocity, in addition to nutrient enrichment (Table 4; Fig. 3). Actually, in Can˜amares River, this green macroalga showed higher % coverage at the upstream sampling site (with lower water velocity and lower nutrient concentrations) than at downstream sampling sites (with higher water velocity and higher nutrient concentrations) (Table 4). On the other hand, the presence of G. densa, P. pectinatus, A. nodiflorum, M. spicatum and R. peltatus at downstream sampling sites, these macrophyte species being absent at upstream sampling sites (Table 4), could also be due to the potential role of propagules from the reservoirs. Although we have no information about the community of submersed macrophytes inhabiting reservoirs in the study areas, several previous studies have shown that reservoirs can act as a source of propagules for aquatic and riparian vegetation, being able to affect positively (but also
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negatively) hydrochory along dammed rivers (Merritt & Wohl, 2002; Merritt et al., 2010; Hyslop & Trowsdale, 2012). Changes in the value of metrics based on the macroinvertebrate community (Table 3; Fig. 2) reflects a substitution of sensitive macroinvertebrates for tolerant ones. For example, plecopterans, trichopterans and some ephemeropterans decreased in abundance, whereas the abundance of tolerant aquatic snails, beetles, dipterans and amphipods increased (Table 5). Similar changes have been previously found in other impounded rivers where the abundance and diversity of submersed macrophytes increased, and the abundance and diversity of benthic macroinvertebrates decreased, downstream from dams (Ward & Stanford, 1979; Petts, 1984; Armitage et al., 1987; Camargo & Voelz, 1998; Ogbeibu & Oribhabor, 2002; Rehn, 2009). Nevertheless, in our study impounded middle reaches (Fig. 1), these changes seemed to be less evident than those found in impounded upper reaches (Casas et al., 2000; Camargo et al., 2005), probably because middle (and also lower) reaches usually have macroinvertebrate taxa that are more tolerant to environmental perturbations (Washington, 1984; Rosenberg & Resh, 1993). Actually, downstream increases in total density and total biomass of the macroinvertebrate community (Table 3) would be caused, in part, for the observed nutrient enrichment downstream from dams: there were positive relations between phosphate concentrations and the abundance of macroinvertebrate scrapers and collector-gatherers (Fig. 4), probably owing to a positive effect of increased nutrient concentrations on periphyton and
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Table 5 Mean (n = 8) relative abundances of macroinvertebrate families at sampling sites of each impounded river Can˜amares River
Bornova River B1 Ancylidae
0.1
Anthomyiidae
–
Aphelocheiridae Baetidae
B2 10.3 –
B3
C1
Sorbe River
C2
C3
S1
S2
S3
0.3
0.6
1.7
2.6
–
7.0
1.4
–
–
–
–
0.2
–
0.8
1.1
–
–
15.6
21.3
25.3
– 9.6
–
–
13.8
3.4
–
–
15.4
32.1
– 5.1
Brachycentridae
0.7
–
–
–
–
–
–
–
–
Caenidae
–
–
–
–
–
–
–
–
1.7
Chironomidae
0.3
3.6
1.4
0.8
4.1
–
3.2
5.5
8.1
Cordulegasteridae
0.3
0.1
–
0.6
0.3
0.1
0.3
0.1
0.2
Dugesiidae
0.8
–
–
–
–
–
–
–
–
Dytiscidae
–
–
–
0.8
–
–
–
–
–
Elmidae
7.5
8.3
5.8
2.3
5.0
8.8
0.6
23.5
20.7
Empididae
4.0
–
0.3
–
–
–
0.1
–
–
Ephemerellidae
4.1
–
0.7
–
–
–
0.5
–
1.4
Ephemeridae
–
–
–
–
Gammaridae Glossiphoniidae
– –
4.2 –
2.7 1.2
Gomphidae
–
–
Gyrinidae
–
– –
–
–
–
–
11.3 –
12.4 –
66.0 0.7
1.4 –
–
–
–
–
0.4
–
–
–
–
–
–
–
–
0.1
Heptageniidae
4.9
–
–
0.2
Hydrobiidae
–
27.4
2.6
63.9
41.2
11.1
22.5
0.1
0.6
0.6
–
0.4
4.5
9.5
4.9
–
–
–
Hydropsychidae Leuctridae
–
0.2
11.6 –
37.4 0.2
0.7
1.2
0.6
–
1.8
3.8
7.5
0.7
6.6
–
5.9
42.9
Limnephilidae
4.0
4.9
0.3
4.4
0.3
0.3
–
–
1.0
Limoniidae
–
–
–
–
–
–
–
–
0.2
Lumbricidae
0.2
0.6
–
0.4
0.5
0.1
–
0.4
–
Lumbriculidae
–
–
0.3
–
–
–
1.1
–
0.8
Perlidae
0.4
0.1
–
0.6
0.1
–
0.2
–
–
Physidae
0.1
1.4
–
–
0.9
0.1
–
0.4
0.7
Planariidae
–
–
–
–
0.1
–
–
11.5
2.1
Planorbidae Polycentropodidae
– –
– 0.2
– 0.2
– –
– –
0.1 –
– –
– 0.4
– –
Rhyacophilidae
–
–
0.2
0.4
0.1
–
0.2
–
–
Sialidae
–
0.1
0.2
0.2
–
–
–
–
–
Simuliidae
1.0
1.7
52.8
2.1
0.3
Sphaeriidae
0.1
0.9
–
0.2
–
–
–
–
–
Tabanidae
0.1
0.1
–
–
–
–
0.1
–
–
27.7
5.2
0.1
3.1
–
–
–
1.6
1.0
–
–
–
–
–
–
0.2
–
Taeniopterygidae Tipulidae
–
–
19.5
6.1
25.0
2
Relative abundances were calculated on the basis of density estimates (individ./m )
perilithon abundance. These findings agree with other findings in rivers and streams experiencing nutrient enrichment (Hellawell, 1986; Rosenberg & Resh,
1993; Camargo et al., 2005; Camargo et al., 2011). Although we have no information about the macroinvertebrate community inhabiting the bottom of
123
176
Fig. 2 Contributions (expressed as %) of macroinvertebrate feeding groups (scrapers, predators, filter-feeders, collectorgatherers and shredders) to the macrobenthic community at
Fig. 3 PCA analyses with submersed macrophytes and water physicochemical parameters at sampling sites: upstream sites are represented by crosses (B1, C1, S1); downstream sites 1 (B2, C2, S2) are represented by circles; and downstream sites 2 (B3, C3, S3) are represented by triangles. Water physicochemical parameters are shown as black arrows, whereas macrophyte taxa are shown as grey arrows
reservoirs in the study areas, we consider that, conversely to the case of submersed macrophytes, the potential of macroinvertebrate propagules from reservoirs contributing to the macroinvertebrate community at downstream sampling sites would be very limited, since macroinvertebrate richness was higher at upstream sampling sites than at their respective downstream sampling sites, with only few differences regarding taxa composition (Tables 3, 5).
123
Hydrobiologia (2014) 728:167–178
upstream and downstream sampling sites. These contributions were calculated on the basis of biomass estimates
Fig. 4 PCA analyses with macroinvertebrate feeding groups and water physicochemical parameters at sampling sites: upstream sites are represented by crosses (B1, C1, S1); downstream sites 1 (B2, C2, S2) are represented by circles; and downstream sites 2 (B3, C3, S3) are represented by triangles. Water physicochemical parameters are shown as black arrows, whereas macroinvertebrate trophic groups are shown as grey arrows
Conclusions Our research has clearly shown that deep-release storage reservoirs, located in middle reaches of impounded rivers, can act as sources of nutrients (phosphate, particularly), causing nutrient enrichment downstream from dams. This abiotic change affected both the macrophyte and macroinvertebrate communities. In the case of submersed macrophytes, total coverage and taxa richness increased downstream, with the species Apium nodiflorum, Groenlandia densa, and Potamogeton pectinatus being only found downstream from dams. In the case of benthic
Hydrobiologia (2014) 728:167–178
macroinvertebrates, total density and total biomass also increased downstream, but taxa richness tended to decrease. Scrapers appeared to be the macroinvertebrate feeding group most favored downstream from dams as a probable consequence of the positive effect of phosphate on periphyton and perilithon abundance. Nutrients would ultimately come from water runoff over agricultural lands and over forests and pastures, being subsequently accumulated in the hypolimnion of reservoirs. We however consider that, if terrestrial vegetation was not completely removed before filling reservoirs, this fact could not contribute currently to the downstream nutrient enrichment owing to the long time passed since dams were constructed and reservoirs developed. Acknowledgments Funds for this research came from the Spanish Ministry of Science and Innovation (research project CGL2011-28585). The University of Alcala´ provided logistical support. Alfonso Benı´tez-Mora was supported by a doctoral fellowship from the Chilean National Commission for Scientific and Technological Research (CONICYT).
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