RIVER RESEARCH AND APPLICATIONS
River Res. Applic. 29: 1090–1099 (2013) Published online 5 July 2012 in Wiley Online Library (wileyonlinelibrary.com) DOI: 10.1002/rra.2595
PATTERNS OF BENTHIC INVERTEBRATE RICHNESS AND DIVERSITY IN THE REGULATED MAGPIE RIVER AND NEIGHBOURING NATURAL RIVERS N. E. JONES* River and Stream Ecology Lab, Ontario Ministry of Natural Resources, Trent University, 2140 East Bank Drive, Peterborough, Ontario, Canada K9J 7B8
ABSTRACT Fluctuating flows common in hydropeaking operations present biota with contrasting and challenging environments. Taxa that require a narrow range of water velocity or are not adapted to withstand sudden changes in discharge will likely be eliminated or competitively disadvantaged under such circumstances, perhaps leading to reduced biodiversity. I investigated the whole river, longitudinal and lateral patterns of benthic invertebrate abundance, Shannon–Wiener diversity, and rarefied taxa density and richness in the hydropeaking Magpie River and 16 neighbouring natural rivers. The Magpie River had greater abundances of benthic invertebrates than natural rivers, particularly near the dam. General differences in benthic community characteristics were largely based on the near absence of Odonata and Plecoptera and an abundance of snails and worms in the Magpie River. Family density, richness and diversity were greater in the regulated Magpie River and unregulated upper Magpie River than found in natural rivers. Longitudinally, family density, diversity and particularly richness increased downstream in the Magpie River. Laterally, diversity did not show any trends with increasing depth along transects, except at near the dam where it decreased sharply with depth, velocity, and an abundance of filter feeding invertebrates. Taxa density did not show any lateral trends in natural rivers, whereas in the Magpie River, it increased with water velocity and depth. The results of this study are contradictory to the general findings of others implying reduced biodiversity below hydropower facilities. Possible explanations are examined and contrasted with other examinations of benthic invertebrate response below hydropeaking dams. © Her Majesty the Queen in Right of Canada 2012. key words: hydropeaking; daily flow fluctuations; richness; diversity; rarefaction; varial zone Received 18 October 2011; Revised 4 May 2012; Accepted 13 June 2012
INTRODUCTION The current and future demand for renewable energy is increasing (ISO, 2007), and this can be seen in the development and modernization of hydropower facilities. There is also a trend for more hydropeaking facilities that can rapidly meet electricity demands during high consumption periods (Scruton et al., 2005). During nonpeak periods, hydropeaking facilities store water in reservoirs, often resulting in low flows below the dam. During peak hours, power is generated and water is released producing much larger flows. These daily fluctuations in flow present biota with contrasting and challenging environments. As noted by Trotzky and Gregory (1974), many regulated rivers are ecologically two different rivers in one: a low flow and a high flow. Taxa that require a narrow range of water velocity or are not adapted to withstand rapid changes in discharge will likely be eliminated or competitively disadvantaged under such circumstances, resulting in reduced biodiversity (Connell, 1978; Munn and Brusven, 1991; Townsend and Hildrew, 1994). *Correspondence to: N. E. Jones, River and Stream Ecology Lab, Ontario Ministry of Natural Resources, Trent University, 2140 East Bank Drive, Peterborough, Ontario, Canada K9J 7B8. E-mail:
[email protected] Reproduced with the permission of the Minister of Ontario Ministry of Natural Resources.
© Her Majesty the Queen in Right of Canada 2012
Maximizing biodiversity and native species richness is often implicitly or explicitly the objective of natural resource policy (May, 1988). The alteration of flow and other corresponding factors (e.g. thermal and sediment regimes) are frequently thought to be the most serious threats to aquatic biodiversity in rivers (Bunn and Arthington, 2002). There are a considerable number of studies and reviews that have examined the effects of flow alteration in river ecosystems. Many of these have examined fish response (e.g. Murchie et al., 2008), and some have examined benthic invertebrate response (e.g. Armitage, 1984; Dewson et al., 2007; Haxton and Findlay, 2008). Few studies have focused on the effects of intermittent hydropeaking on benthic invertebrates (but see Garcia de Jalon et al., 1994; Valentin et al., 1995; Cereghino and Lavandier, 1998; Cereghino et al., 1997; Almodóvar and Nicola, 1999; Cortes et al., 2002; Fuller et al., 2010). The response of invertebrates downstream of hydropower dams is not universal and appears to depend on a variety of interacting factors, for example, design and operation of the dam and environmental context. Some studies reported decreases in benthic abundance and/or biomass below a peaking facility (Moog, 1993; Garcia de Jalon et al., 1994; Cereghino and Lavandier, 1998; Cereghino et al., 1997), other studies reported increases (Fuller et al., 2010), and still others
BIODIVERSITY OF BENTHIC INVERTEBRATES IN REGULATED AND NATURAL RIVERS
reported no difference (Valentin et al., 1995; Cortes et al., 2002; Almodóvar and Nicola, 1999). The diversity of taxa found below peaking facilities also varied with some research suggesting a decrease in diversity or simplification of the community (Garcia de Jalon et al., 1994; Valentin et al., 1995), whereas others found no changes (Almodóvar and Nicola, 1999). Unfortunately, the terms richness and diversity are often poorly defined and used interchangeably in most of these studies. In many cases, rarefaction is not used, and species richness is reported when species density (taxa per unit area) was actually measured (Gotelli and Colwell, 2001). Gotelli and Colwell (2001) wrote on the typical issues in quantifying biodiversity in the measurement and comparison of species richness, noting that ecologists have not always appreciated the influence of abundance, system productivity, and sampling effort on richness measures and comparisons. Without the standardization of measures, meaningful comparisons are not truly valid (Vinson and Hawkins, 1998; Gotelli and Colwell, 2001). The Magpie River is the focus of a long-term ecosystembased study on the effect of river regulation and ramping rate restrictions (Smokorowski et al., 2010). Large daily changes in flow on the Magpie River translate into changes in wetted width and a large varial zone (Jones, 2011). Jones (2011) showed that the presence of longitudinal and lateral gradients in invertebrate abundance and the community composition were related to the upstream reservoir and fluctuating flows in the Magpie River. He noted that without careful consideration of these gradient effects, conclusions based on monitoring programs and experiments could be flawed. In this study, I investigate patterns of benthic invertebrate abundance, diversity and rarefied taxa density and richness in the regulated Magpie River and 16 neighbouring natural rivers. I examine patterns in diversity measures among rivers, longitudinally within rivers, and laterally from the shore to deeper waters along sampled transects. I hypothesize that changes in flow, temperature and sediment regimes associated with peaking (Clarke et al., 2008) reduce richness and diversity of benthic invertebrate communities. I used a rarefaction technique (Colwell, 2005) to generate expected community measures using natural reference rivers and an upstream–downstream (control impact) study design.
METHODS Study rivers All study rivers are located on the Boreal Shield along the north shore of Lake Superior, Canada (Figure 1). The Magpie River, Wawa, Ontario (48 400 N; 84 4400 W), has a flow regime altered by the Steephill Falls waterpower facility (WPF). The Magpie River drains an area of 1930 km2 down to Lake Superior and has a mean annual flow of 27 m3 s1 © Her Majesty the Queen in Right of Canada 2012
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Figure 1. Study region along the north shore of Lake Superior
showing sampling rivers and their watersheds. See Table 1 for characteristics of the watersheds
observed during a 51-year record preceding the initial operation of the WPF in 1991. The WPF began operation in 1990 and was required to include minimum flow and ramping rate restrictions as part of its operating requirements. The peaking facility draws water from a depth of 10 m in the 30-m deep reservoir and supplies a metalimnetic (cool water) draw that supports a recreational brook trout fishery. The maximum and the minimum discharges through the turbines are 44 and 7.5 m3 s1, respectively. Before the fall of 2004, ramping was restricted to 1 m3 s1 h1 from 10 October to 15 November and 2 m3 s1 h1 from 16 November until spring freshet (early May). From May until early October, restrictions included an increase or decrease of 25% of the previous hour’s flow. After the summer of 2004, ramping rates were deregulated, and unrestricted flow alterations were allowed: increases in flow from the minimum to the maximum, or vice versa, in less than 1 h became possible. During freshet, additional water is spilled if reservoir levels exceed the rule curve for reservoir level. In dryer springs, water is not spilled. The resultant flow regime consists of daily fluctuations between the minimum and the maximum flows, in addition to any spillage over the dam. This contrasts sharply with the steady daily flows of natural rivers in the region. In the summer, regulated flows peak during daylight hours and drop at night. During some weekends, only the baseflow is discharged. This daily fluctuation in flow creates extensive varial areas where the substrate is exposed and dried in sunlight for hours to days between rewetting flows. The outflow of the dam flows slowly through a 3-km long lentic segment before dropping in gradient and increasing in velocity for the following 25 km. This deep lake-like portion was not sampled. The Magpie River drainage has teardrop-shaped ‘fume-kill’ area extending northeastwards from a sinter plant near the town of Wawa. Sulfur dioxide and metals emissions increased steadily from 1939 to 1973 and then were reduced River Res. Applic. 29: 1090–1099 (2013) DOI: 10.1002/rra
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until the plant closed in 1998 (Greenaway et al., 2012).The fume kill area is 80 km2 or approximately 4% of the watershed area. Despite this disturbance, there has been remarkable chemical and biological recovery and recovery rates may surpass those in many Sudbury lakes (Greenaway et al., 2012). This fume-kill area, once devoid of vegetation, now supports abundant grasses, herbs and sparse stands of trees. Riparian conditions are noticeably different downstream of the hydropower dam for 8 km and may affect the results and their interpretation of data collected at transects one and two in this study. In addition to the regulated Magpie River, 16 natural rivers with little development were studied (Figure 1; Table 1) to provide an understanding of the natural range of variability in diversity patterns and establish benchmarks against which differences of the Magpie River could be compared. Aside from differences in drainage basin area, most rivers have similar habitat characteristics including substrate composition (e.g. cobbles) and water chemistry. The fish communities in rivers along the north shore of Lake Superior are similar and fairly simple (e.g. longnose dace, sculpin, brook trout). Samples were collected in mid-July of each year during 2004, 2005, 2006, 2007 and 2009. The Magpie River and the natural Batchawana River were sampled each year at multiple transects. The Magpie River had three transects spaced 3, 5 and 8 km downstream from Steephill Falls WPF. Similarly, three transects were spaced 3 km apart in the natural Batchawana
River, whereas the other 16 natural rivers had only one transect that was sampled once. The focus on the Batchawana stems from a companion study that only used the Batchawana River as a control (Smokorowski et al., 2010) The upper Magpie River was sampled in 2010 and is located 40 km upstream from Steephill Falls dam (Figure 1). The upper Magpie River represents a natural unregulated river and was used as an internal control for comparison with the regulated Magpie River. Natural rivers along the north shore of Lake Superior experience a predictable perennial flow regime. Upstream lakes on river networks are generally absent or far upstream from sampling areas. Waters are slightly acidic and clear with some humic staining. Most rivers are used by migratory salmonid species for spawning, rearing and opportunistic feeding in the spring and fall. Many of the rivers become too warm and dry in July–August to sustain more than juvenile salmonids in small thermal refugia. Rivers run through a mixture of alluvial and colluvial geomorphology, with frequent outcroppings of bedrock and expansive cobble point bars more typical of western North America systems. Field sampling Benthic invertebrates were sampled to a substrate depth of 10 cm using a 0.09-m2 Surber sampler fitted with a 500-mm mesh and immediately preserved in 70% ethanol. Eight to ten Surber samples were equally spaced laterally along the
Table I. Landscape-scale characteristics of the sampled watersheds Name
Code
Area km2
Mean Elevation [masl]
MAP [mm]
GDD >5 C
Baseflow Index
Conductivity mS
Arrow River Current River Wolf River Jackpine River Cypress River Gravel River Steel River White River White Gravel River Pukaskwa River Dog River Upper Magpie River Magpie Rivera Old Woman River Agawa River Batchawana River Chippewa River
AR CU WO JA CY GR ST WH WG PU DG UpMag MAG OW AG BAT CH
783 639 727 285 164 627 1202 5220 108 1099 1229 1324 1773 225 1168 1226 718
472 460 389 443 448 438 378 401 404 431 434 394 388 354 458 437 417
801 724 711 829 837 839 821 861 879 924 946 897 911 1027 1065 1072 1070
1224 1128 1135 1007 991 986 1002 1030 930 937 985 1066 1067 1071 1082 1173 1206
0.53 0.58 0.55 0.57 0.59 0.60 0.60 0.59 0.59 0.62 0.62 0.60 0.61 0.61 0.59 0.60 0.61
100 60 150 129 70 90 129 110 80 80 70 135 100 136 133 120 121
1089 108 5220
420 354 472
895 711 1072
1060 930 1224
0.59 0.53 0.62
107 60 150
Map ID 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17
Mean Min Max
Watershed area, mean elevation of the watershed, mean annual precipitation (MAP), growing degree days >5 C (GDD), baseflow index developed by Neff et al. (2005) used to represent groundwater potential and water conductivity. a Magpie River regulated flow. © Her Majesty the Queen in Right of Canada 2012
River Res. Applic. 29: 1090–1099 (2013) DOI: 10.1002/rra
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Figure 2. Schematic of a transect across a river. Numbers indicate sampling points where benthic invertebrates were collected with a Surber. In the Magpie River, the varial zone (wet–dry) is indicated as the area of river that is frequently wetted and dried between 7.5 and 44.0 m3 s1 (from Jones, 2011)
transects, leading from the wetted shallows to progressively deeper waters near the thalweg (Figure 2). In the Magpie River, sampling took place for 2 days: once during high flows when the varial zone could be sampled and once during low flows when the deeper permanently wetted streambed could be sampled. Transects were selected in riffle areas that gradually increased in water depth (mean = 0.2 m, SD = 0.10), velocity (mean = 0.38 m s1, SD = 0.25) and that showed large changes in wetted width in relation to changes in stage height. Water depth (cm), velocity (m s1; measured using a Swoffer 2100; Swoffer Instruments Inc. Seattle, WA, USA) and substrate size composition were recorded at each sampling point. Substrates were classified as clay and silt (256 mm). Effort was made to sample areas with similar substrate size composition to reduce variability from other potentially confounding factors. In the laboratory, invertebrates were identified to genus, except Nematoda, Turbellaria and terrestrial invertebrates, which were typically identified to Family. Each taxon was classified as terrestrial, lentic or lotic based on their habitat preferences (Merritt et al., 2008). Analyses To compare invertebrate abundance, I used one-sample t-tests where the natural rivers form a distribution for comparison with a single mean estimated for the Magpie River, upper Magpie River and transects along Magpie and Batchawana rivers. One-sample t-tests were used because while numerous natural rivers were sampled, only one regulated Magpie River was sampled. To understand how rarefied family density varies with river size, I regressed family density in relation to drainage basin area using simple linear regression. The upstream reservoir on the Magpie River was known to provide and support at least 12 additional lentic taxa such as Daphnia and Hydra, which were also very abundant (Jones, 2011). These taxa were removed from analyzes because they are typical of lake environments and they would have heavily influenced richness and diversity measures. SigmaPlot 12 was used for analyses. I used © Her Majesty the Queen in Right of Canada 2012
the Shapiro–Wilk test to examine data for normality and the Levene median test for homogeneity of variances. For all statistical tests, I used an a = 0.05 as a critical level of significance. I used correspondence analysis (CA) to ordinate associations among invertebrate abundance from the Magpie River and natural rivers, which illustrate general differences in invertebrate communities at the order level. The matrix of river stream-years and taxa were (log(w + 1) transformed to down weight the influence of both dominant and rare taxa (Keller et al., 2002). Dragonflies were the rarest group but still found at 65% of streams. Data for the CA were analyzed using Biplot (Lipkovich and Smith, 2001). This indirect gradient analysis procedure is robust and consistent (Jackson, 1993) and provides information about the similarity between rivers based on the proportional abundance of taxa present in each community (Legendre and Legendre, 1998). To compare taxa density, richness and Shannon–Wiener diversity among rivers, transects and samples along transects, rarefaction curves were generated using EstimateS software, Storrs, CT, USA (Colwell, 2005) using samplebased family level data. EstimateS is a computer program that calculates multiple rarefaction curves and statistical estimates of true richness using Monte Carlo resampling techniques. The nonparametric Chao-1 estimator was used to generate smooth rarefaction curves and associated 95% confidence intervals because it often performs better than other estimators (Chao, 1984; Walther and Moore, 2005). Each estimate was based on means calculated from 1,000 randomizations with replacement. The final value for the averaged, random-order species accumulation curve with replacement is generally less the total number of observed species, but the advantage is that the estimates of variance remains meaningful at the right-hand end of the species accumulation curve and can thus be used to compare data sets. Taxa density (the number of families in a sampled area) and Shannon–Wiener diversity were determined for a set number of samples, and thus a sampling area (square metres) common to each comparison. This scaling removes the effect of differences in the abundance of individuals between rivers. River Res. Applic. 29: 1090–1099 (2013) DOI: 10.1002/rra
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Family richness was determined for a set number of individuals common to each comparison (as per Gotelli and Colwell, 2001). This contrasts with taxa density in that sampled area is not the unit of comparison but rather the number of individuals captured is used to scale comparisons. Estimates of taxa density, richness and diversity were calculated for individual all rivers, transects and samples along transects. One-sample t-tests were used to test for differences between mean diversity measures from the regulated Magpie River and values of diversity estimated from natural rivers. To make visual comparisons of lateral trends in richness and diversity measures from across rivers and years, invertebrate diversity and taxa density data were scaled to 1.0, where the largest value within an invertebrate group from a transect was used as the denominator. Scaling the data allows rivers of different productivity to be compared by bringing them to a common scale. Simple linear regression was used to examine the significance and sign of the relationship between invertebrate abundance and sampling location along the transect (distance from shore; see Jones, 2011 for similar analyses on abundance data).
RESULTS In the regulated Magpie River, transect one and the upper Magpie River had greater abundances of benthic invertebrates than natural rivers, including the Batchawana (OS t-test, P < 0.001; Figure 3). Much of the abundance in the regulated Magpie River stems from the high production near the dam, but this longitudinal pattern was not found in the natural Batchawana River (Figure 3). By transect three
Figure 3. Abundance of invertebrates during 2004, 2005, 2006, 2007 and 2009 in the regulated Magpie River (Mag) and natural rivers, including the upper Magpie River (UpMag) and Batchawana River (Batch). Transects (T1, T2 and T3) on the Magpie River were 3, 5 and 8 km downstream from the hydropower dam. Transects of the Batchawana were similarly spaced. Whiskers on the plot are standard errors and are merely there to illustrate relative variability in the estimate. Asterisks indicate significant differences between 16 natural rivers and the mean values from other grouped transects in one-sample t-tests © Her Majesty the Queen in Right of Canada 2012
Figure 4. Relationship between rarefied family density (families per unit area) found in study rivers and the respective drainage areas upstream for the sampling area
invertebrate densities in the Magpie River were more similar to those found in natural rivers. As expected, estimated taxa density increased as a function of drainage basin area (N = 16, F = 8.02, P < 0.013, R2 = 0.36; Figure 4). The similarity, or dissimilarity, among rivers based on the proportional abundance of taxa present in each river can clearly be seen in the results of our CA (Figure 5). Axis 1
Figure 5. CA on abundance of the major taxonomic groups found in the Magpie River and 16 natural rivers for each transect sampled. Natural rivers are indicated by solid black dots, the Magpie River by circles. River codes are listed in Table 1. T1, T2 and T3 suffixes refer to longitudinally spaced sampling transects. Transects sampled over multiple years were averaged before analysis River Res. Applic. 29: 1090–1099 (2013) DOI: 10.1002/rra
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accounted for 40% of the total variance, and axis 2 accounted for 30%. Axis 1 split the transects sampled in natural rivers from the regulated Magpie River based largely on the presence and absence of Odonata, Plecoptera, snails and Trichoptera. Axis 2 eigenvalues were heavily weighted by Odonata, worms and snails. Natural rivers tended to have low abundances of Odonata and Plecoptera, whereas the regulated Magpie River had almost none. Natural rivers also tended to have few Coleoptera. Samples from the regulated Magpie River had an abundance of worm-like organisms (e.g. Enchytraeidae, Lumbricidae, Naididae) and snails (Basommatophora). There were no leeches (Erpobdellidae), Peaclams (Pisidiidaein), brushlegged mayflies (Isonychiidae), dragonflies (Aeshnidae, Gomphidae) and Limnephilidae caddisflies in the Magpie River, which are fairly common in natural rivers. Absent in natural rivers were Lymnaeidae snails and Unionicolidae and Pionidae water mites. Natural rivers had a greater abundance of Chironominae whereas the Magpie River had greater numbers of Orthocladiinae. Rarefied taxa richness and density and Shannon–Wiener diversity indices were significantly greater in the upper Magpie River than that found in natural rivers (OSH t-test, P < 0.001; Figure 6). Similarly, taxa richness, diversity and density were significantly greater in the regulated Magpie River than that found in natural rivers (OSH t-test, P < 0.001). The upper Magpie River was not significantly different than the regulated Magpie River for rarefied taxa richness and density, and diversity based on 95% confidence intervals. The relationship between taxa density and the estimated number of invertebrates for a standard six samples (per transect) asymptote after approximately 4000–5000 individuals or at approximately 35–40 taxa (Figure 7). It appears that taxa density decreases at very high benthic densities in the
Figure 6. Estimated taxa density and richness and Shannon–Wiener
diversity for the regulated Magpie River and natural rivers, including the upper Magpie River. Taxa density and diversity were generated for six samples and richness was generated using 2 000 individuals using family level data. Standard errors are shown for the 16 natural rivers © Her Majesty the Queen in Right of Canada 2012
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Figure 7. Relationship between rarefied taxa density (family) and the associated estimate number of individuals sampled. Highlighted are transect one samples for the years 2004, 2005, 2006, 2007 and 2009 in the Magpie River close to the dam where invertebrates abundance and water velocities were highest
Magpie River, particularly for samples collected closest to the dam at transect one. Taxa density and Shannon–Wiener diversity index were not significantly different among transects one, two, and three in the regulated Magpie River (Figure 8). However, there was some indication that these measures increased from transect one downstream to transect three. Taxa richness at transect one was significantly lower than that calculated for transect three (Figure 8). Similar longitudinal trends were not found in the Batchawana River (data not shown).
Figure 8. Estimated taxa density and richness and Shannon–Wiener diversity for transects one, two and three in the regulated Magpie River, 3, 5 and 8 km downstream from the hydropower dam. Data collected from 2004, 2005, 2006, 2007 and 2009. Rarefied taxa density and diversity were generated for 40 samples and richness was generated using 15 000 individuals. Whiskers on the plots are 95% confidence intervals River Res. Applic. 29: 1090–1099 (2013) DOI: 10.1002/rra
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Figure 9. Patterns of invertebrate diversity and taxa density along sampling transects for the regulated Magpie River and natural rivers including the upper Magpie River (plots A and B) and for transects one, two and three on the Magpie River that were 3, 5 and 8 km downstream from the hydropower dam (plots C and D). Standard errors and whiskers are provided for natural rivers. To make visual comparisons of lateral trends data were scaled to 1
Shannon–Wiener diversity along transects in the Magpie River and natural rivers did not show any lateral trends with increasing depth of samples (Figure 9a). Overall diversity was greater in the upper Magpie River (paired t-test, N = 8, P < 0.001) but not between natural rivers and the regulated Magpie River (paired t-test, N = 8, P = 0.35). Taxa density did not show any lateral trends along transects in natural rivers but increased with water velocity and depth in the Magpie River (R2 = 0.74, P < 0.001, N = 8; Figure 9B). Diversity at transect one showed a significant negative trend with depth and velocity (R2 = 0.80, P < 0.001, N = 10; Figure 9C). Taxa density increased from the shoreline to deeper and faster areas of the Magpie River (Figure 9D).
DISCUSSION This study describes patterns of family level richness, density and diversity in the hydropeaking Magpie River and in nearby natural rivers. Many authors have noted the negative effects of river regulation on benthic invertebrates (e.g. Ward and Stanford, 1979; Armitage, 1984; Dewson et al., 2007; Haxton and Findlay, 2008); however, family level density and richness in the regulated Magpie River were greater than that observed in neighbouring natural © Her Majesty the Queen in Right of Canada 2012
rivers. These findings suggest that the Magpie River has not been seriously affected by regulation and contrast with the findings of Moog (1993), Garcia de Jalon et al. (1994), Cereghino and Lavandier (1998) and Cereghino et al. (1997), who found significant decreases in the abundance and/or biomass of benthic invertebrates below peaking facilities. Only Valentin et al. (1995), Cortes et al. (2002) and Almodóvar and Nicola (1999) found no effect, and Fuller et al. (2010) noted an increase in abundance/biomass. The diversity of taxa found below peaking facilities has been shown to decrease (Garcia de Jalon et al., 1994; Valentin et al., 1995), whereas others found no changes (Almodóvar and Nicola, 1999). A lack of explicit definitions of richness measures and their estimation make interstudy comparisons difficult. This issue aside, past studies on hydropeaking effects may have one or more other shortcomings, for example, (i) poor site selection (e.g. directly below dam) and few sites to determine longitudinal pattern or reference condition (e.g. Munn and Brusven, 1991; Valentin et al., 1995), (ii) insufficient time after impact to allow changes and colonization to occur (Munn and Brusven, 1991; Garcia de Jalon et al., 1994) and (iii) few details in regarding spatial arrangement of samples, particularly laterally with respect to being in wet or dry zones (e.g. Gore, 1977; Garcia de Jalon et al., 1994; River Res. Applic. 29: 1090–1099 (2013) DOI: 10.1002/rra
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Valentin et al., 1995). Many of these studies sampled immediately below newly installed dams and only sampled at one or two sites to compare with one control site above the dam (e.g. Garcia de Jalon et al., 1994; Valentin et al., 1995). Further, others commenced sampling only a year or two after the dam and reservoir were built (Garcia de Jalon et al., 1994). Such a short period does not allow new taxa to colonize the river, whereas one or two sampling sites does not allow the zone of influence/effect to be determined. Studies that sample longitudinally can identify how impacts change with distance from the dam (Gore, 1977). We should likely expect changes in benthic communities below newly created dams. In many cases, the reservoir will be transporting newly flooded materials, changing the thermal regime (e.g. warm to cold water communities) and changing the sources of energy for invertebrates (Ward and Stanford, 1983). It will likely take many years for a river and invertebrate community to approach a new equilibrium (Clarke et al., 2008; Murchie et al., 2008 Box II). In the present study, sampling occurred 24 years after the hydropower dam was built and likely allowed adequate time for many changes to occur including colonization of biota (Thienemann, 1954, cited in Hynes, 1970; Mackay, 1992). Benthos sampling was distributed longitudinally along the river so that the affected zone could be determined. Ultimately this is a question of what is the appropriate amount of time for recovery and what is the true spatial area of impact (Minns, 2006)? What if the composition of the community has changed while biodiversity measures and function have not? More emphasis needs to be placed on adaptive management, long-term cycles, and the use of regulated rivers in an experimental manner to learn about fundamental aspects of ecology and how to best manage hydropower developments. Transect one on the Magpie River had the highest invertebrate abundance (individuals m2) and, perhaps as a consequence, it also had the lowest rarefied taxa richness, density, and diversity compared with transects further downstream. This finding is similar to other studies that sampled just downstream of a dam (Gore, 1977; Garcia de Jalon et al., 1994; Valentin et al., 1995; but not so severe as Munn and Brusven, 1991). Munn and Brusven (1991) found few taxa but very high abundance below the Dworshak Reservoir on the Clearwater River, Idaho. They suggested that low habitat diversity, fluctuating water levels, altered thermal regime and possibly food supply were causes for the simplified invertebrate community. Finding simplified invertebrate communities immediately below hydropower dams is common in the literature (e.g. Gore, 1977; Valentin et al., 1995; Haxton and Findlay, 2008). In the Magpie River, estimated family density appeared to asymptote at this transect or perhaps even decline at high abundances. Members of this simplified community included those specialized in filtering large quantities of © Her Majesty the Queen in Right of Canada 2012
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seston leaving the reservoir upstream (e.g. Trichoptera (Ceratopsyche morosa) and Simulidae; Jones, 2011). Water velocities were also the highest (~0.9 m s1) at transect one.The low richness at transect one is likely do to a combination of environmental tolerance, ecological specialization, and perhaps competitive exclusion (Grime, 1974; Townsend and Hildrew, 1994; Vinson and Hawkins, 1998). The abundance of invertebrates decreased downstream of the dam from 100,000 at transect one down to 50,000 and 30,000 at transects two and three, respectively. Downstream of the dam the abundance of filter feeding invertebrates likely decreased as drifting seston density declined by settling to the river bed or by selective consumption (Richardson and Mackay, 1991; Jones, 2010). It is unknown if or how the flume kill area influenced the longitudinal gradients found. Transect three was 3.0.CO;2-X Chao A. 1984. Non-parametric estimation of the number of classes in a population. Scandinavian Journal of Statistics 11: 265–270. Clarke KD, Pratt TC, Randall RG, Scruton DA, Smokorowski KE. 2008. Validation of the flow management pathway: effects of altered flow on fish habitat and fishes downstream from a hydropower dam. Canadian Technical Report of Fisheries and Aquatic Sciences 2784: vi + 111 p. Colwell RK. 2005. EstimateS: Statistical estimation of species richness and shared species for samples. Version 7.5. User’s Guide and application published at: http://viceroy.eeb.uconn.edu/estimates Connell JH. 1978. Diversity in tropical rain forest and coral reefs. Science 199: 1302–1310. Corbet PS. 1999. Dragonflies: Behavior and Ecology of Odonata. Cornell University Press: Ithaca, NY; 12–427. River Res. Applic. 29: 1090–1099 (2013) DOI: 10.1002/rra
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