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The effects of streamside forest harvesting on aquatic macroinvertebrates and rainbow trout (Oncorhynchus… Mackenzie, Kirsten Dawn 2005

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THE EFFECTS OF STREAMSIDE FOREST HARVESTING ON AQUATIC MACROINVERTEBRATES AND RAINBOW TROUT (ONCORHYNCHUS MYKISS) DIET IN THE CENTRAL INTERIOR OF BRITISH COLUMBIA, CANADA By KIRSTEN DAWN MACKENZIE B.SC. UNIVERSITY OF NORTHERN BRITISH COLUMBIA, 2000 A THESIS SUBMITTED IN PARTIAL FULFILMENT OF THE REQUIREMENTS FOR THE DEGREE OF MASTER OF SCIENCE IN THE FACULTY OF GRADUATE STUDIES (FORESTRY) THE UNIVERSITY OF BRITISH COLUMBIA APRIL 2005 © Kirsten Dawn MacKenzie, 2005 ABSTRACT Timber harvesting has the potential to alter stream ecosystems in a variety of ways, impacting stream habitat, food resources, and ambient conditions for a variety of stream biota. While ample attention has been focused on salmonid-bearing streams in the coastal areas of British Columbia, comparatively little research has been applied to the interior regions of the province where climate, topography, and forest practices differ wildly from the wet, mountainous west coast. This thesis investigates the effect of streamside forest harvesting on the benthic and drift macroinvertebrate communities, and on the diets of stream-resident rainbow trout in three small, lake-headed streams in the central interior of B.C. Benthic invertebrate biomass and abundance, community composition, and functional feeding guild composition did not exhibit post-logging changes in the two treatment streams relative to the control. Similarly, the biomass, abundance, community composition, and functional feeding guild composition of the drift invertebrate community did not show marked differences between logged and unlogged streams. Rather, patterns within both of these communities tended to reflect natural temperature differences evident among the streams. Rainbow trout in the two logged streams exhibited significantly lower condition (weight for a given length) than trout in the unlogged control stream. Trout in the control stream ate a greater biomass and abundance of invertebrates than trout in the logged streams. Trout in the control stream also tended to eat larger invertebrates than fish in the logged streams; however, this did not reflect differences in the size of prey items available in the drift. Two families of invertebrates tended to dominate the biomass and abundance of the stomach contents of all fish: Ephemeroptera (mayflies) and Diptera (true flies). Neither the size, abundance, nor biomass of these primary prey organisms were affected by the logging treatment. The unique temperature regimes coupled with density dependence associated with these small lake-headed streams were likely the most important factors in regulating invertebrate communities and fish growth in these small streams. TABLE OF CONTENTS Abstract ii Table of Contents iii List of Tables v List of Figures vi Acknowledgements viii CHAPTER 1: The effects of forest harvesting on benthic invertebrate standing crop and invertebrate drift in streams of the B.C. central interior 1 1.1 Introduction;., 1 1.2 Materials and Methods 4 1.2.1 Study area 4 1.2.2 Benthic invertebrate sampling 6 1.2.3 Drift invertebrate sampling 6 1.2.4 Data analysis 7 1.3 Results 9 1.3.1 Benthic invertebrates 9 1.3.2 Drift invertebrates 11 2.1 Discussion 13 CHAPTER 2: The effects of forest harvesting on rainbow trout condition, diet, and food resources in the central interior of British Columbia 39 2.1 Introduction 39 2.2 Materials and methods 41 2.2.1 Study area 41 2.2.2 Rainbow trout condition and growth 42 2.2.3 Trout stomach contents 42 2.2.4 Data analysis 44 2.3 Results 45 2.3.1 Rainbow trout condition and growth 45 2.3.2 Trout diet 45 2.3.3 Trout primary prey items 47 2.4 Discussion 49 REFERENCES CITED 67 iv LIST OF TABLES TABLE PAGE 1.1 Physical characteristics of the study streams 20 1.2 List of invertebrate taxa collected in benthic samples and their abundance in the study streams in 1997 and 2001 21 1.3 List of invertebrate taxa collected in drift samples and their abundance in the study streams in 2004 22 2.1 List of invertebrate taxa collected in rainbow trout stomachs and their abundance in July and August 2001 23 v LIST OF FIGURES FIGURE PAGE 1.1 Map showing the locations of the three study streams 29 1.2 Mean (±1 standard error) biomass and abundance of invertebrates sampled from the benthos before logging in 1997 and after logging 2001 30 1.3 Canonical variates scores based on the biomass and abundance of organisms within benthic invertebrate taxa 31 1.4 Canonical variates scores based on the biomass and abundance of organisms within functional feeding guilds 32 1.5 Mean drift invertebrate biomass per cubic meter of water by stream, site, and month 33 1.6 Mean total drift invertebrate abundance in the logged streams 118/16 and 118/48, and unlogged control Hip Creek in July and August 2001 34 1.7 Canonical variates scores based on the biomass of organisms within invertebrate drift taxa by site within stream and month within stream 35 1.8 Canonical variates scores based on the abundance of organisms within invertebrate drift taxa compared between sites within streams in July and August 2001 36 1.9 Canonical variates scores based on the biomass of organisms within drift functional feeding guilds 37 1.10 Canonical variates scores based on the abundance of organisms within drift functional feeding guilds compared between sites within streams in July and August 2001.. 38 2.1 Graph of the length-weight regression (relative condition) for resident rainbow trout in the study streams 60 2.2 Mean biomass and abundance of invertebrates in the stomach contents of rainbow trout 61 2.3 Canonical variates scores on the biomass and abundance of invertebrate taxa in the stomach contents of rainbow trout 62 2.4 Mean individual invertebrate weight (mg) in the drift and gut contents of trout in the study streams 63 2.5 Mean individual dry mass (mg) of the three primary prey groups in the stomachs of rainbow trout 64 2.6 Mean (±1 standard error) total biomass and abundance of the total gut contents, and of the three primary prey families in the drift samples and stomachs of rainbow trout from the three study streams 65 2.7 Mean (±1 standard error) total biomass of Ephemeroptera in the gut contents of rainbow trout in the study streams 66 vii ACKNOWLEDGEMENTS I would like to express my sincere gratitude to my supervisor, Dr. Scott Hinch and to my supervisory committee members, Dr. John Richardson and Mr. Eric Parkinson for their support in this project. Also to my many field and lab assistants whom I could never have lived without: Leslie Chamberlist, Shirley Fuchs, Trevor Nowak, Mark Whelly, Johanna Ledezma, Vicki Maloney, Stephanie Topp, Shannon Luttin, Christina Andersen, and Celia Chung. Special thanks to Greg Pearson, Peter Baird, Ed Staub, and the staff at CANFOR Fort St. James Woodlands Division for their financial and logistical support, for looking out for "us girls", and for not once flagging my truck. Also, a heartfelt thank-you to my family, friends, and lab-mates for their words of encouragement, constructive criticism, and for the fish. Financial support was provided by grants to Dr. Hinch by Forest Renewal British Columbia and the Forest Innovation Investment, and by a Science Council of BC GREAT Scholarship to KDM. Finally, I dedicate my thesis to Eric Mellina, without whom I wouldn't be the person I am today. Thank you for teaching me the value of loyalty and hard work, for being a pillar of support, and for encouraging me when I needed it the most. And thanks for not putting me in the hospital..."Doctor". Chapter 1: The effects of forest harvesting on benthic invertebrate standing crop and invertebrate drift in sub-boreal streams of the B.C. central interior INTRODUCTION Streamside timber harvesting has varied effects on salmonids (Oncorhynchus spp.) and their habitats. Deleterious effects include increased sediment and debris loads (Platts et al. 1989; Burns 1972), increased nutrient input from overland runoff (Hansmann and Phinney 1973; Murphy and Hall 1981; Noel et al. 1986; Gregory et al. 1987), increased solar radiation leading to increased temperature (Hetrick et al. 1998b, Platts and Nelson 1989), and decreased habitat stability and complexity (Chamberlin et al. 1991). However, the increased nutrients and solar radiation can cause an increase in stream primary productivity (Murphy and Hall 1981; Murphy et al. 1981; Shortreed and Stockner 1983; Murphy et al. 1986; Beschta et al. 1987; Anderson 1992; Stone and Wallace 1998), which can lead to increased benthic invertebrate abundance (Hawkins et al. 1982; Stone and Wallace 1998), biomass, and diversity (Newbold et al. 1980; Murphy et al. 1981; Stone and Wallace 1998). The effects of increased invertebrate production may be mirrored in fish growth and production (Murphy and Hall 1981; Hetrick et al. 1998a), provided that stream temperatures remain below the lethal limits of the species. The effects of environmental disturbance in small stream ecosystems may be carried through the food chain in a "bottom-up" manner; that is, the lowest trophic levels are the first to respond to disturbance. In the case of forest harvesting, increased light resulting from the decreased canopy cover has been shown to increase stream primary production (Hansmann and Phinney 1973; Murphy and Hall 1981; Shortreed and Stockner 1983; Hetrick et al. 1998b; Fuchs et al. 2003). Aquatic invertebrates, in particular grazers and scrapers (Merritt and Cummins 1996), which feed on biofilm, may experience enhanced growth and survival as a result (Gurtz and Wallace 1984; Anderson 1992; Hetrick et al. 1998a; Fuchs et al. 2003). 1 Leaf litter and other terrestrial-derived detritus are another primary food source for stream invertebrates (Anderson and Sedell 1979; Vannote et al. 1980; Hall et al. 1981, Scrivener 1987). In most systems, streamside timber removal results in a net loss of allochthonous organic matter inputs to streams (Murphy and Hall 1981). Leaf-litter exclusion studies have shown that such loss can result in a decline in the abundance of shredders (Baer et al. 2001; Siler et al. 2001), which can in turn influence the abundance of invertebrate detrivores and their vertebrate predators (Wallace et al. 1997). However, loss of organic matter is only temporary in some locales as red alder (Alnus rubra) and other understory plants begin to dominate the riparian zone and contribute organic matter to streams. In southeast Alaska, streams flowing through young-growth alder forests exported significantly more invertebrates to downstream reaches than young-growth conifer forests (Piccolo and Wipfli 2002). Most benthic invertebrates drift at some point in their lifecycle. Drifting may be a mechanism by which populations of benthic invertebrates can regulate their density, access food, or escape predators. Some benthic invertebrates may also get involved in the drift when they pupate, or when they return to the stream to lay eggs as adults (for example, Diptera and Ephemeroptera; see review by Brittain and Eikeland 1988). From a fish predation perspective, drift is important as a means by which benthic invertebrates become available to fish feeding in the water column. Benthic invertebrates are a main food source for many stream fish (see Bryan and Larkin 1972; Scott and Crossman 1973; Allan 1981; Dedual and Collier 1995). Drift composition is often highly correlated with the composition of the benthic community (Brittain and Eikeland 1988); however, some invertebrate families drift more than others (Waters 1965; Townsend and Hildrew 1976). Larvae and nymphs of aquatic insects are generally the most common invertebrates in the drift, but in certain systems zooplankton, molluscs, and leeches may occur in high numbers (Rader 1997; Brittain and Eikeland 1988). Because drift is largely made up of benthic invertebrates, conditions that affect the benthos may have an impact on the composition of the drift. In these situations, drift may mirror benthic communities with regard to 2 composition and abundance. Alternatively, drift communities may differ substantially from benthic communities since benthic macroinvertebrates drift in response to a variety of factors that differ among species and streams (Rader 1997). Most studies investigating effects of logging on invertebrate and fish populations have occurred in coastal ecosystems (e.g. Newbold et al. 1980; Murphy and Hall 1981; Hawkins et al. 1982; Culp and Davies 1983; Murphy et al. 1986; Anderson 1992), which are characterised by mountainous or hilly terrain, high winter storm frequency, and high precipitation (Heede 1984). However, much of the future timber supply in the northern hemisphere is expected to come from boreal and sub-boreal forests (Bryant et al. 1997). In these regions, moderate hillslope gradients dominate, and the climate is characterised by low winter storm frequency and low precipitation (Heede 1984). Thus, streams and their biota in these regions may react differently to logging as a result of differing climate, soil types, forest types, and runoff regimes. One of the few studies to examine effects of timber harvesting on stream characteristics and benthic invertebrate communities in sub-boreal streams was Fuchs et al. (2003). They found streams with recently logged riparian on both banks (less than 5 years prior to the study) had higher chlorophyll a biomass and higher total benthic macroinvertebrate biomass than older logged streams (timber harvest over 20 years old) and those with old-growth riparian. The authors speculated that higher algal biomass in recently logged streams either led to faster benthic invertebrate growth rates, resulting in larger individuals and hence higher total biomass, or that smaller sized taxa may have been replaced by larger-sized taxa, also resulting in a higher biomass. Contrary to their expectations given the higher algal biomass, recently logged streams showed the lowest density and biomass of invertebrates in the scraper guild compared to the other treatments. The authors suggested that low scraper biomass was observed because its biomass peaked early in the season before sampling took place. This was supported by the high densities of predator and parasite guilds, biota which had probably taken 3 advantage of the early, high abundance of scrapers. Melody (2000) surveyed several small sub-boreal streams and found no differences between logged and unlogged sections in periphyton, detritus, macroinvertebrate density, and biomass. However, most of her streams were subjected to forest harvesting on only one bank, which likely muted the logging effects and may account for differences with the Fuchs et al. (2003) results. Nonetheless, these studies, and most of those conducted in coastal systems, used comparative survey approaches, which can reveal strong correlations. However, subtle correlations can be obscured by large internal variation, and it is difficult to assess cause and effect with these approaches. I used a before-after-control-impact (BACI) design to experimentally test the hypotheses that riparian timber harvest from sub-boreal streams resulted in an overall increase in the biomass and abundance of benthic macroinvertebrates. I also tested the specific hypothesis that scraper abundance and biomass in logged streams peaked in early summer, as suggested by Fuchs et al. (2003). The BACI contrast focused on invertebrates captured in the benthos; however, I compared post-logging drift invertebrate communities with those from the benthos. Drift may better reflect the availability of invertebrate prey to fish (Brittain and Eikeland 1988), so understanding how or if drift differs from benthos in logged streams is important for understanding how fish may be affected by logging. MATERIALS AND METHODS Study area I examined three small, lake-headed streams (118/16, 118/48, and Hip Creek, Figure 1.1) located within the sub-boreal spruce biogeoclimatic zone of north-central BC (Farley 1979). This region is dominated by glaciolacustrine and sandy glaciofluvial soils (MacDonald et al. 1992; Larkin et al. 1998). Annual precipitation averages 50cm, falling primarily as snow in winter, with seasonal peaks in late spring and early autumn (MacDonald et al. 1992). The 4 hydrograph is dominated by spring meltwater runoff (Cheong et al. 1995); however, a smaller peak occurs in autumn due to increased rainfall (Beschta et al. 1987). The streams were matched as closely as possible with respect to their physical characteristics (see details in Mellina 2002). All of the streams were headed by small, shallow lakes, although Hip Creek had a wetland complex situated between the lake and the study site. Details of the physical and thermal characteristics of each stream are presented in Table 1.1. Temperature data represents averages for the period June 1 - August 31, 2001. On each stream, a study section 400-600 m in length was delineated by upstream (US) and downstream (DS) sampling sites. No tributaries occurred within the study sections. Fish species composition in all streams was dominated by rainbow trout (Oncorhynchus mykiss), although small numbers of suckers (Catostomus spp.), lake chub (Couesius plumbeus), prickly sculpins (Cottus asper), and burbot (Lota lota) were present (Mellina 2002). Each stream was studied one year before logging (1997) and 3 years after logging (2001). Logging was conducted by Canadian Forest Products in February 1998. The riparian and portions of upland around streams 118/16 and 118/48 (hereafter termed the "treatment streams") were clear-cut, and all merchantable timber (>15 cm diameter at breast height (DBH) for lodgepole pine (Pinus contorta) and >20 cm DBH for spruce (Picea spp.) and sub-alpine fir (Abies lasiocarpa)) was removed. A 5 m buffer from which heavy machinery was excluded was maintained around the treatment streams to preserve the soil quality and bank stability. Invertebrate sampling sites were located within the upstream and downstream boundaries of the clear-cuts in the treatment streams. Clearcut size was 40 ha for 118/16 and 36 ha for 118/48 (Mellina 2002). Hip Creek served as a control stream and was not subjected to any forest harvesting treatment. 5 Twenty benthic samples per stream per year were collected with a Surber sampler (0.09 m 2 quadrant area, 350 pm mesh, substrate was disturbed to a depth of 10 cm) during late October 1997 (pre-logging), and early November 2001 (post-logging). Samples were collected from riffles approximately 20 m apart throughout the study sections, and washed in a 350 pm sieve before being preserved in 10% buffered formalin. Due to time and resource limitations, only 10 to 16 samples per stream were examined and used in analyses. Insects were identified to the lowest possible taxonomic level (usually genus) (Stewart and Stark 1988; Wiggins 1996; Merritt and Cummins 1996), while other organisms were identified to phyla (e.g. Nematoda) or class (e.g. Hirudinea). A list of taxa collected and their abundance in the samples is presented in Table 1.2. Each taxon was also assigned to a functional feeding guild using information provided in Merritt and Cummins (1996). Samples were transferred to filter paper, dried overnight at 80°C and weighed to the nearest 0.001 mg. The biomass of each taxonomic group in each sample was estimated from dry weight. For each sample, total biomass and abundance of invertebrates was also calculated. Drift nets (diameter 153 mm, mesh 350 u.m) were anchored in the streams at the upstream and downstream boundaries of the study areas. They were emptied six times each month over a 24-hour period (approximately every 4 hours) in July and August, 2001 and samples were preserved in 10% formalin. Each time the drift nets were set, water depth and velocity were measured at the mouth of the drift net. This allowed for the calculation of the volume of the water that passed through the net during the sampling period, which was then used to standardise invertebrate biomass and abundance measures. All drift samples were sorted, identified to lowest possible taxonomic level (usually genus) (Merritt and Cummins 1996; Stewart and Stark 1988; Wiggins 1996; Borror et al. 1989; Bland and Jaques 1978; Chu 1978), and dried overnight at 80°C. Dried samples were weighed to the nearest 0.001 mg. A list of taxa collected and their abundance in the drift samples is presented in Table 1.3. Taxa were also assigned to a functional feeding guild using information from Merritt and Cummins (1996). 6 Terrestrial insects were not assigned to a feeding guild, but to a 'Terrestrial" category. For each sample, total biomass and abundance was also calculated. Invertebrate families were used for analyses of community composition. This simplified the analyses by reducing the number of variables. Too many variables in the analysis may have resulted in false findings of significant differences among classes. Rare taxa (i.e. where a family was present in less than 50% of the samples) were combined into one category (other aquatic) to reduce the number of zeroes in the dataset. Biomass and abundance data were logio(>f/+ 1) transformed to meet the assumptions of normality required for parametric tests. I used two-way analysis of variance (ANOVA) to examine the effects of stream and year on total abundance and biomass of benthic invertebrates. Post-hoc comparisons were made using Bonferroni t-tests. Two-way multivariate analysis of variance (MANOVA) with stream and year as classes was used to examine differences in invertebrate community composition and functional feeding guild structure. Analyses were conducted separately on abundance and biomass datasets. If MANOVA identified that year, stream, or year-by-stream interaction had an effect on community or functional guild composition, canonical variates analysis (CVA) was performed to describe the differences between multivariate centroids. The CVA scores along the two most significant canonical axes (as determined by the proportion of the total variation explained by the axis) were plotted to visualize differences between streams and years. Variables were considered to contribute to the total variation explained by the axis if the total standardized canonical coefficient was greater than 0.32 or less than -0.32 (McGarigal et al. 2000). For all analyses, a significant interaction effect would indicate that a change in one of our invertebrate variables (e.g. abundance, biomass, community composition, or feeding guild structure) from 1997 to 2001 differed among streams, which indicates that the logging treatment had an effect on the invertebrate community composition. 7 Two-way analysis of variance (ANOVA) was used to analyze for differences between streams in the total biomass and abundance of invertebrates in the drift, as well as the biomass and abundance of invertebrates in each taxonomical group and functional feeding guild. However, since pre- and post- logging samples were not available, samples from the upstream sites on each stream were treated as unlogged since the water and organisms entering the site were not affected by the logging treatment located downstream. Samples taken from the downstream sites on 118/48 and 118/16 were considered to come from logged sites since the water and organisms passing through the area had been subjected to the full effect of the logging treatment upstream. I looked for significant interaction effects between stream and site to determine if the total biomass and abundance of invertebrates in taxonomic groups and feeding guilds reacted differently between upstream and downstream sites in the logged streams versus the unlogged control. If MANOVA identified that site, stream, month, or any interactions among these classes were significant, CVA was performed to describe the differences between multivariate centroids. The scores were then plotted along the two most significant canonical axes (determined by the proportion of the total variation explained by the axis) to visualize the results among sites and streams. Variables were considered to contribute to the total variation explained by the axis if the total standardized canonical coefficient was greater than 0.32 or less than -0.32 (McGarigal et al. 2000). A significant interaction effect would indicate that a change in one of our invertebrate variables (e.g. abundance, biomass, community composition, or feeding guild structure) from the upstream site to the downstream site differed among streams or months, which indicates that the logging treatment may have had an effect on the invertebrate community composition. For all analyses, results were considered significant if the probability of a false significant result was less than 0.05 (i.e. 95% confidence). SAS software (SAS Institute Inc., 1988) was employed for all analyses. 8 RESULTS Total biomass (mg-m2) of benthic invertebrates was not affected by the year*stream interaction (ANOVA, P = 0.99), indicating that the logging treatment did not affect biomass. Total biomass did not differ between streams (P = 0.60); however, it was higher in all streams in 2001 compared to 1997 (ANOVA, P < 0.01; Figure 1.2a). I detected no year-by-stream interaction for the total abundance (numberm'2) of benthic invertebrates (ANOVA, P=0.66, Figure 1.2b). Total abundance was higher in 2001 compared to 1997 (ANOVA, P = 0.03), and both years, abundance was higher in Hip Creek than stream 118/16 (P = 0.03). Abundance in stream 118/48 was intermediate between Hip Creek and 118/16 but did not differ from either. A list of invertebrate taxa and their abundance in the surber samples is presented in Table 1.2. Biomass of benthic invertebrate taxa differed among streams (MANOVA, P < 0.01). The first canonical axis explained 82% of the variation in the data, and revealed that stream 118/16 differed from stream 118/48 and Hip Creek by having a higher biomass of Diptera larvae, Pelecypoda, and organisms in the 'other' category (including snails, leeches, and rare insect taxa) (CVA, P < 0.01, Figure 1.3). The second canonical axis explained 18% of the variation and indicated that stream 118/48 had a higher biomass of Coleoptera larvae, Pelecypoda, and Trichoptera larvae, and a lower biomass of Acarina, and Ephemeroptera nymphs than Hip Creek (CVA, P < 0.01). Abundance of invertebrates within benthic invertebrate taxa differed among streams (MANOVA, P = 0.02). The first canonical axis explained 81% of the variation and indicated that stream 118/16 contained a greater abundance of Diptera, Oligochaeta, and Trichoptera, whereas stream 118/48 and Hip Creek contained a greater abundance of Acarina, Coleoptera larvae, Collembola, Ephemeroptera, Megaloptera, and Plecoptera (CVA, P < 0.01 for all comparisons, Figure 1.4). The second canonical axis explained the remaining 19% of the variation and represented a gradient between streams abundant in Ephemeroptera, Odonata, and Oligochaeta, and streams abundant in Coleoptera larvae and adults, Trichoptera, and 9 organisms in the 'other' category. Hip Creek and stream 118/16 scored low on this axis, while stream 118/48 scored slightly higher. While the effect of year within stream on the abundance of invertebrates in benthic taxa was not significant (MANOVA, P = 0.07), trends in the data revealed further differences between streams. The first canonical axis explained 48% of the variation, while the second canonical axis explained 24%. Stream 118/48 and Hip Creek scored higher on both axes in 2001 relative to 1997, while stream 118/16 scored lower on the first axis and higher on the second axis in 2001. In the first two streams, this represents an increase in the abundance on Acarina, Ephemeroptera, Plecoptera, Trichoptera, and other aquatic organisms. Conversely, stream 118/16 experienced an increase in the abundance of Coleoptera larvae and adults, Diptera, Trichoptera, and other aquatic invertebrates. The biomass of invertebrates within functional feeding guilds differed among streams (MANOVA, P < 0.01). The first canonical axis explained 77% of the variation, and indicated that stream 118/16 differed from both stream 118/48 (CVA; P < 0.01) and Hip Creek (CVA; P = 0.03) by having a greater biomass of scrapers and parasites, and a lower biomass of shredders and collectors (Figure 1.5). Stream 118/48 did not differ from Hip Creek (CVA; P = 0.28). The second canonical axis explained the remaining 23% of the variation, and represented a gradient between streams with high biomass of collector-gatherers and predators to streams with high biomass of shredders, scrapers, and collector-filterers. Abundance of invertebrates in functional feeding guilds also differed among streams (MANOVA, P < 0.01). The first canonical axis explained 88% of the variation, and indicated that stream 118/16 differed from stream 118/48 and Hip Creek by having a more collector-gatherers, shredders, and parasites/piercers and fewer predators and scrapers (CVA, P < 0.01 for both comparisons, Figure 1.6). Stream 118/48 did not differ from Hip Creek (CVA, P = 0.18). The second canonical axis explained the remaining 12% of the variation and represented a gradient between streams with abundant collector-filterers and shredders to streams with abundant collector-gatherers and parasites/piercers. Total invertebrate drift biomass per cubic meter of water was higher at the upstream sites than the downstream sites (ANOVA, P = 0.02, Figure 1.7b), and lower in August than in July (P < 0.01, Figure 1.7c) in all streams. In addition, there were significant differences in invertebrate biomass among streams. Stream 118/16 had the highest drift biomass, followed by stream 118/48 and Hip Creek; however, only stream 118/16 and Hip Creek were significantly different (P = 0.02, Figure 1.7a). Total abundance of drift differed between months within streams (ANOVA, P < 0.01). Drift was higher in stream 118/16 and Hip Creek in July relative to August (P < 0.05 for both comparisons), whereas stream 118/48 did not differ between months (Figure 1.8). A list of the taxa encountered in the drift, and their abundance is presented in Table 1.3 MANOVA indicated that there was an effect of site within stream on the biomass of invertebrate families (P < 0.01). Canonical variates analysis revealed that the upstream and downstream sites differed on stream 118/16 and in Hip Creek, but not in stream 118/48. In stream 118/16, the composition of invertebrate taxa in the drift shifted from high biomass of Araneae (mainly mites), Diptera pupae, and Ostracoda at the upstream site to a higher biomass of Ephemeroptera and Plecoptera nymphs at the downstream site (CVA, P < 0.01, Figure 1.9). A similar, but non-significant pattern was observed in stream 118/48 (CVA, P = 0.13). In Hip Creek, from the US to the DS sites there was a shift from Ephemeroptera nymphs, terrestrial insects, and Collembola in the drift at the upstream site to a higher biomass of Diptera pupae and Gastropoda at the downstream site (CVA, P = 0.01). Biomass of drift taxa was affected by the interaction between month and stream (MANOVA, P < 0.01). Composition differed between months in stream 118/16 and Hip Creek (CVA, P < 0.05 11 for both comparisons), but not in stream 118/48 (CVA, P = 0.23). During July, drift samples from stream 118/16 contained a high biomass of Collembola, Ostracods, Hymenoptera adults, and other terrestrial insects (Figure 1.10). In August, samples from the same stream contained a higher biomass of Diptera adults and larvae, Araneae, and Gastropoda. A similar but non-significant pattern was observed in stream 118/48. Hip Creek in July was intermediate on axis 1 and high on axis 2 indicating a high biomass of Diptera larvae. In August, Hip Creek scored slightly lower on both axes, indicating a shift towards a higher biomass of Plecoptera nymphs, Araneae, and Gastropoda. Analysis of the invertebrate community composition based on the abundance of invertebrates in drift taxa produced different results from those observed with biomass. A three-way interaction between site, stream, and month indicated that the relationship between upstream invertebrate composition and downstream invertebrate composition differed between streams and months (Figure 1.11). In July, drift composition in stream 118/16 differed from that in stream 118/48 and Hip Creek (CVA, P < 0.01 for both comparisons). Drift composition on stream 118/16 also differed between sites in July (CVA, P < 0.01). In July, the upstream site on stream 118/16 contained abundant Coleoptera and Diptera larvae, Araneae, Gastropoda, and other terrestrial insects, whereas the downstream site contained abundant Hymenoptera adults, Araneae, and Gastropoda. Drift composition did not differ between sites on the other two streams (all sites, P>0.05), or between the streams themselves (all streams, P>0.05). In August, there were no differences between any of the sites or streams (CVA, P>0.05). There were effects of site within stream, and month within stream on the biomass of invertebrates within functional feeding guilds (MANOVA, P < 0.05 for both classes). Feeding guild composition differed between US and DS sites on stream 118/16 (CVA, P < 0.01), but not in the other streams (118/48: P = 0.18; Hip: P = 0.99). The first canonical axis explained 59% of the variation and indicated that the upstream site of stream 118/16 contained a higher biomass 12 of collectors and terrestrial insects, whereas the downstream site contained a higher biomass of scrapers and shredders (Figure 1.12). The second canonical axis explained 30% of the variation. Drift biomass within feeding guilds also differed between July and August in stream 118/16 (CVA, P < 0.01), but not in the other two streams (118/48: P = 0.17, Hip: P = 0.45). The first canonical axis explained 67% of the variation, and indicated that in July, stream 118/16 had a significantly higher biomass of terrestrial insects compared to August, when it contained a higher biomass of predators, collector-gatherers, and scrapers. The second canonical axis explained 16% of the variation in the dataset. Analysis of the abundance of invertebrates in functional feeding guilds revealed an effect of site within stream between months similar to that found with the abundance of invertebrates within taxa (MANOVA, P < 0.01). The first canonical axis explained 47% of the variation and indicated that in July, there were a large number of scrapers at the upstream site of stream 118/16 compared to the downstream site, which had abundant collector-gatherers and shredders (Figure 1.13). In August, 118/16 upstream contained high numbers of collector-filterers and terrestrial insects, compared to the downstream site, which contained relatively high numbers of collector-gatherers and parasites (CVA, P < 0.01). In stream 118/48 and Hip Creek, there were no differences between sites in either month (CVA; P>0.05 for all comparisons). The second canonical axis explained 27% of the variation in the dataset. DISCUSSION My results indicate that among-stream patterns in benthic invertebrate community were not related to the logging treatment. Past studies which have examined streamside logging often have reported an increase in benthic invertebrate abundance and/or biomass (Newbold et al. 1980; Murphy and Hall 1981; Murphy et al. 1981; Hawkins et al. 1982; Fuchs et al. 2003), and 13 changes in invertebrate community structure (Gregory et al. 1987, Hawkins et al. 1982). Logging-induced changes were mainly attributed to increased solar energy leading to enhanced primary production and concomitant shifts in food type and availability. In my study, the vegetative canopy covered over 70% of total stream area prior to logging. Because it was reduced by -50% immediately following logging on the two treatment streams (Mellina 2002), I had anticipated that biomass and abundance of benthic invertebrates would increase in the treatment streams relative to the control (as in Fuchs et al. 2003), and that scraper biomass would increase due to increased periphyton growth (as in Hawkins and Sedell 1980). One possible reason I found no changes attributable to enhanced primary production in that my post-logging observations were made three years after logging, so the potential increase in solar radiation reaching the treatment streams may not have been extreme enough due to rapid regeneration of deciduous riparian trees and shrubs. Although photosynthetically active radiation was not measured, all of the streams (control and treatment) had a canopy that shaded 40 to 60% of their surfaces. Furthermore, deciduous vegetation grows into a denser canopy than the original coniferous one in years immediately following riparian harvest (Murphy and Meehan 1991). It is possible that I may have found some differences in benthic invertebrates attributable to logging-related changes in primary production had I examined treatment streams in the year immediately following logging. Some investigators have suggested that increases in allochthonous inputs immediately following streamside logging (i.e. from leaf litter or slash) can lead to an increase in abundance or biomass of invertebrates from functional feeding guilds that process leaf litter and detritus (i.e. shredders and collector-gatherers), and a decrease in periphyton feeding guilds (i.e. scrapers) (Hawkins and Sedell 1980; Cummins et al. 1989). Furthermore, studies in the central interior of British Columbia have found that litter input, not light, regulates invertebrate communities in experimental mesocosms (Melody and Richardson 2004). In my study, I found no changes in feeding guild composition or invertebrate abundance or biomass that were 14 consistent with a post-logging related explanation. It is possible that changes to the invertebrate community occurred immediately following the logging treatment when streamside vegetation was less abundant than it was before logging, or three years after logging. It is reasonable to expect that with the removal of streamside vegetation, invertebrates such as collectors-gatherers and shredders decreased in abundance due to the loss of litter inputs. Since this study was conducted three years following logging, regeneration of the streamside vegetation may have also led to rehabilitation of these invertebrate communities. Piccolo and Wipfli (2002) suggest that canopy type also affects the amount and type of invertebrates in small streams. Deciduous vegetation such as red alder and twinberry are abundant, even under the coniferous canopy in this region; thus, there may not have been as dramatic a shift in litter types entering the stream as I had expected. The most likely reason that I observed no post-logging changes in benthic invertebrate measures was that stream temperature did not change much as a result of logging. Invertebrate community patterns and total abundance or biomass are directly affected by changes in stream temperature (Merritt and Cummins 1996). In headwater streams, riparian logging commonly causes larger increases in stream temperature during summer months (Hicks et al. 1991; Macdonald et al. 2003). Mellina et al. (2002) studied the responses of logging-induced stream temperature patterns in the same streams that I investigated. The study streams are headed by small lakes and thus were naturally warmer prior to logging, and they tended to cool as they travelled downstream due to the influence of cold groundwater inputs. Mellina et al. (2002) found that the initial temperature of the water, combined with the groundwater entering streams such as ours was the primary determinant of stream temperature as it passed through clearcuts in the sub-boreal region of British Columbia. Streams 118/16 and 118/48 were naturally warmer than Hip Creek during the summer months (Mellina et al. 2002), and post-logging changes to temperature patterns in the treatment streams were minor compared to those generally observed in headwater streams in sub-boreal areas (Mellina 15 unpublished data). The among-stream differences in summer temperatures, which were unrelated to logging, seemed to coincide with the differences that I found among stream invertebrate communities. My results suggest a division between streams with different ambient conditions, rather than a division between logged and unlogged streams. Specifically, trends in benthic community composition were more similar between stream 118/48 and Hip Creek (the cooler streams) than in stream 118/16 (the warmer stream). Both 118/48 and Hip Creek contained a higher biomass of Acarina (mites), Coleoptera larvae, and Plecoptera nymphs compared to stream 118/16, which contained a higher biomass of Diptera larvae, Pelecypoda (clams), and other aquatic organisms. Plecoptera in particular tend to be associated with clean, cool, running water, while Diptera in general are adapted to a much wider range of conditions (Merritt and Cummins 1996; Cox and Rutherford 2000). Within streams, differences between years also seemed related to differences in stream temperature. Temperatures were somewhat cooler in 2001 (Table 1.1), and stream 118/48 and Hip Creek both experienced an increase in the abundance and biomass of invertebrates that prefer cooler water, such as Plecoptera (stoneflies) and Trichoptera (caddisflies). In summary, natural differences among streams and years in temperature appeared to play a more important role in regulating benthic invertebrate community patterns than did any habitat changes caused by streamside logging. As with the benthos, my results did not indicate that logging had a strong or consistent effect on the drift community in my streams. However, trends in the abundance and biomass of invertebrate drift did not mirror the trends in benthic invertebrate abundance and biomass. Stream 118/16 contained the greatest number of drifting invertebrates, but also the lowest biomass, indicating that the invertebrates were smaller than those in other streams. Sweeney and Vannote (1978) found that aquatic insects metamorphosed into adults at a smaller size in warmer streams than in cooler ones. This further supports my assertion that natural temperature differences among streams play a larger role in regulating invertebrate communities than any habitat changes caused by streamside logging. 16 It is difficult to elucidate patterns in community and functional feeding guild composition in the drift. Invertebrates drift for a variety of reasons, including predator avoidance, emergence, and to seek out new food sources (Brittain and Eikelahd 1988); therefore, it may be difficult to elucidate the specific effects that forest harvesting may have on the drift community. In all analyses, stream 118/16 stood out from the other streams, but the variables explaining the variation between 118/16 and the other streams differed with each analysis. In stream 118/16,1 found a shift in biomass from taxa that are more tolerant of high temperatures at the upstream site, to more temperature-sensitive taxa at the downstream site. This may reflect the cooling trend in water temperature, from the upstream maximum of up to 30°C to the downstream maximum of 25°C (Mellina et al. 2002). In Hip Creek, where stream temperatures were consistently low throughout the summer, trends in community composition did not reflect any particular water quality indicator, and in stream 118/48, where water temperatures were moderate, but rarely exceeded optimal thermal ranges for most organisms, there were no downstream changes to community composition at all. Shifts in streamside vegetation from coniferous to deciduous have been shown to influence the input of terrestrial organisms to streams (Wipfli 1997; Allan et al. 2003). The biomass of invertebrates in the drift both in the functional feeding guilds and in the various taxa indicated that in July, stream 118/16 was dominated by inputs of terrestrial invertebrates (mainly Hymenoptera and other taxa such as mayflies and stoneflies). However, since these patterns do not occur in either the control stream or the other treatment stream, it is unlikely that the logging treatment is causing an increase in the biomass of terrestrial insects. It is possible that these organisms entered the drift while metamorphosing into adults or while laying eggs on the water surface, the timing of which may differ widely among streams (Brittain and Eikeland 1988). 17 Functional feeding guild abundance patterns in the drift also stood out in stream 118/16, differing among months within streams within years in each analysis. Scrapers, parasites, and collector-gatherers were responsible for most of the variation between sites. Collector-gatherers may be more prone to drifting since they feed on detritus, which tends to be unevenly distributed throughout the stream (Hildebrand 1974; Merritt and Cummins 1996). Parasites may also be more local in distribution (for instance, existing in high densities where host densities are higher), and their contribution to the functional feeding guild structure may have been skewed by their local abundance. Terrestrial invertebrate abundance did not stand out among these data as they did in the biomass results, a finding that is supported by Wipfli (1997), who found that terrestrial invertebrates tended to be slightly larger than their aquatic counterparts, accounting for a larger proportion of the biomass than of their abundance in the drift. The data presented in this study, contrary to previous research in this region (for example, Fuchs et al. 2003), do not reflect any trends that can be attributed to the logging treatment. This study focused exclusively on three streams headed by small lakes, whereas previous research in this region has either focused entirely on headwater streams, or has not distinguished between headwater and lake-origin streams (Fuchs et al. 2003, Melody 2000). Lake-headed streams are common in sub-boreal and boreal areas and make up approximately 30% of the fish-bearing streams in the region where this research was conducted (MacKenzie, unpublished data). They represent unique systems with regard to temperature regime (Mellina et al. 2002) but are poorly studied in terms of their ecology and potential impacts from forestry. It is possible that the combination of extreme summer temperatures and low flows in these streams has had a unique effect on the invertebrate community. Most of the within-stream differences existed only among the drift invertebrates - communities that are highly variable both among and within streams (Brittain and Eikeland 1988) - while among-stream differences seemed to reflect differences in temperature patterns rather than logging treatment. 18 The type of logging conducted around small streams may also be important to the outcome of a study such as this. Our logging treatments were conducted 'carefully', with no destabilization of the stream banks, no removal of large woody debris, and limited retention of streamside vegetation and non-merchantable timber, whereas the streams studied by Fuchs et al. (2003) were clear-cut to the banks with no retention of streamside vegetation. The sub-boreal location meant that these streams were also not affected by the steep topography, heavy precipitation, and unstable soils that predominate in coastal areas, where much of the research into streamside forest harvesting has occurred. The careful logging methods used in my study, combined with the unique thermal characteristics of sub-boreal, lake-headed streams likely diminished the negative impacts of high temperatures and sediment inputs that are commonly reported in studies conducted in coastal or sub-boreal headwater systems (e.g. Meehan 1991) Sub-boreal streams provide habitat to vast numbers of fish and invertebrate species, many of which are economically and culturally valuable, yet they remain largely unstudied. Increasing global demand for resources means that resource extraction will likely increase in the coming years; thus, it is important to understand how activities such as logging affect these ecosystems. Limited sub-boreal research on headwater streams shows that there is a potential risk of damage to invertebrate communities (Fuchs et al. 2003); however, my results suggest that if logging is conducted carefully around lake-headed streams in these areas, there may be little impact on resident invertebrate communities. These results are consistent with those reported by Mellina (2002), who found that rainbow trout populations in my study streams, and stream physical habitat, were also not affected by the logging treatments. 19 Table 1.1. Physical characteristics of the three study streams. Data are from Mellina (2002). Values for stream and hillslope gradients, and bankfull width and height were based on five measures taken along the study section and averaged. Temperature data loggers were positioned at the upstream (US) and downstream (DS) boundaries of the study sites. Temperature data represent averages, based on daily mean and maximum values for the period June 1 - August 31, 2001. Stream Length Bankfull Bankfull Aspect Stream Hillslope US Mean US DS Mean DS of study width height gradient gradient Temperature Maximum Temperature Maximum section (m) (cm) (%) (%) (°C) Temperature (°C) Temperature (m) (°C) (°C) 118/16 372 1.7 29.5 SE 1.8 19.8 15.13 19.87 13.65 16.68 118/48 607 1.7 35.2 SW 1.9 26.7 14.83 16.28 13.52 15.04 HIP 430 1 26.7 N 3.1 32.7 11-9 13.16 11.43 12.63 20 Table 1.2. List of invertebrate taxa and their abundances collected in surber samples in the study streams in 1997 and 2001. Larva nymph = (n), pupa = (p), adult = (ad). Taxa 118/16 118/48 HIP Grand Class/Order Family Genus 1997 2001 1997 2001 1997 2001 Total Acarina Hydracarina 1 1 Acarina 4 2 4 3 2 21 Amphipoda Gammaridae Gammarus 1 1 Arachnida 3 3 11 Cladocera Daphniidae Daphnia 1 1 Cladocera 3 5 Coleoptera Elmidae Ampumixis 3 2 5 Coleoptera Elmidae Cleptelmis 1 7 6 4 1 30 Coleoptera Elmidae Heterlimnius 1 1 3 1 4 11 Coleoptera Elmidae (ad) 4 1 5 Collembola 1 1 3 Copepoda 6 6 5 29 Copepoda/Calanoida 1 1 2 Diptera Ceratopogonidae Probezzia 2 6 1 8 1 4 28 Diptera Ceratopogonidae 1 1 Diptera Chironomiidae Chironomiinae 7 8 8 6 6 6 52 Diptera Chironomiidae Orthocladiinae 5 7 8 7 4 6 47 Diptera Chironomiidae Tanypodiinae 7 7 7 4 4 3 43 Diptera Chironomiidae Tanytarsini 1 1 Diptera Dixiidae Dixa 1 Diptera Dixiidae Meringodixa 1 Diptera Empididae Chelifera 1 1 3 3 9 Diptera Empididae Oreogeton 2 2 Diptera Empididae 1 Diptera Psychodidae Pericoma/ Telmatoscopus 3 4 8 8 7 4 45 Diptera Simuliidae Prosimulium 1 1 Diptera Simuliidae Simulium 2 4 4 3 1 18 Diptera Simuliidae 1 1 Diptera Tipuliidae Dicranota 5 6 7 4 5 1 33 Diptera Tipuliidae Hexatoma 2 1 3 6 Diptera Tipuliidae Tipula 1 2 Diptera (1) 1 1 Ephemeroptera Ameletiidae Ameletus 1 1 Ephemeroptera Baetidae Baetis 1 5 7 6 4 35 Ephemeroptera Caenidae Caenis 1 1 Ephemeroptera Ephemerellidae Drunella grandis 4 4 4 3 4 3 22 Ephemeroptera Ephemerellidae Drunella spp. 1 1 2 4 Ephemeroptera Ephemerellidae 3 4 8 Ephemeroptera Heptageniidae Cinygma 2 2 2 3 23 Ephemeroptera Heptageniidae Cinygmula 3 1 2 6 Ephemeroptera Heptageniidae Epeorus 3 5 1 5 22 Ephemeroptera Heptageniidae Heptagenia 1 1 3 Ephemeroptera Heptageniidae Rhithrogena 1 1 Ephemeroptera Heptageniidae 8 7 6 5 39 Ephemeroptera Leptophlebidae Choroterpes 4 1 1 6 Ephemeroptera Leptophlebidae Paraleptophlebia 3 1 5 3 3 22 Ephemeroptera Leptophlebidae 1 1 1 6 Ephemeroptera Siphlonuridae 1 1 Ephemeroptera (n) 1 Gastropoda 3 1 6 1 14 Hirudinea 2 1 Megaloptera Sialidae Sialis 1 2 2 1 7 Odonata Corduliidae Epitheca 1 1 Odonata Corduliidae Somatochlora 1 1 Oligochaeta Naididae (ad) 1 1 Oligochaeta 6 8 6 7 2 3 41 Ostracoda 1 5 2 8 Pelecypoda 7 8 5 9 2 41 Plecoptera Capniidae Paracapnia 1 4 5 6 4 3 25 Plecoptera Capniidae 2 1 4 3 3 1 22 Plecoptera Chloroperlidae Neaviperla 5 5 2 17 Plecoptera Chloroperlidae Paraperla 1 2 3 Plecoptera Chloroperlidae Sweltsa 1 1 8 7 5 4 38 Plecoptera Chloroperlidae 2 1 1 6 Plecoptera Leuctridae Despaxia 1 5 1 2 12 Plecoptera Leuctridae 2 7 2 4 5 22 Plecoptera Nemouridae Malenka 1 1 2 Plecoptera Nemouridae Zapada 1 4 4 8 5 4 37 Plecoptera Perlidae Hesperoperla 5 Plecoptera Perlidae 2 1 10 Plecoptera Perlodidae Isoperla 1 1 Plecoptera Perlodidae Megarcys 1 1 Plecoptera Perlodidae Skwala 1 2 2 1 2 1 10 Plecoptera Perlodidae 1 2 4 7 1 15 Trichoptera Brachycentridae Micrasema 3 3 2 8 Trichoptera Glossosomatidae Glossosoma 2 8 7 4 4 31 Trichoptera Hydropsychidae Himalopsyche 1 1 Trichoptera Hydropsychidae Hydropsyche 3 1 3 16 Trichoptera Hydropsychidae 1 Trichoptera Hydroptilidae Oxyethira 4 4 Trichoptera Lepidostomatidae Lepidostoma 4 1 2 2 1 1 14 Trichoptera Limnephilidae Chyranda 1 1 Trichoptera Limnephilidae Hesperophylax 2 3 Trichoptera Limnephilidae Hydatophylax 1 1 Trichoptera Limnephilidae Limnephilius 1 1 Trichoptera Limnephilidae 1 4 1 7 Trichoptera Phryganeidae Ptilostomis 1 1 Trichoptera Polycentropodidae Polycentropus 1 2 3 Trichoptera Rhyacophilidae Rhyacophilia 2 1 8 5 6 4 37 Trichoptera Rhyacophilidae (P) 1 1 Trichoptera 1 1 3 Trichoptera (1) 2 1 2 1 6 Trichoptera (p) 1 1 Grand Total 97 111 184 184 130 120 1101 Table 1.3. List of taxa and their abundance in the drift samples of the study streams. (I) = larvae, (n) = nymphs, (p) = pupae, (imm) immature, (ad) = adults. Taxa 118/16 118/48 HIP Grand Class/Order Family Genus July August July August July August Total Acarina 428 147 202 47 143 54 1513 Amphipoda Gammaridae Gammarus 11 5 1 1 22 Arachnida 11 5 4 6 8 13 60 Cladocera 174 36 135 4 5 878 Coleoptera Copeletus (ad) 4 1 2 6 14 Coleoptera Amphizoidae Amphizoa (I) 16 1 5 1 25 Coleoptera Dytiscidae Celina (ad) 2 8 3 9 1 23 Coleoptera Dytiscidae Uvarus (ad) 1 1 Coleoptera Dytiscidae (ad) 2 2 Coleoptera Elmidae Cleptelmis (I) 1 2 5 2 1 11 Coleoptera Elmidae (ad) 4 3 9 Coleoptera Georyssidae Georyssus (I) 3 7 26 Coleoptera Gyrinidae Gyrinus (ad) 1 1 2 Coleoptera Gyrinidae Gyrinus (I) 1 1 Coleoptera Haliplidae Haliplus (ad) 3 3 Coleoptera Hydrophilidae (I) 1 1 2 Coleoptera Hygrobiidae (I) 1 1 Coleoptera Staphylinidae (ad) Staphylinidae (ad) 5 5 3 2 42 48 159 Coleoptera Staphylinidae (I) Staphylinidae (I) 2 1 2 2 9 Coleoptera (ad) 26 1 10 13 11 9 104 Coleoptera (I) 1 2 3 Collembola 14 26 10 17 89 18 236 Collembola Sminthuridae 2 4 23 7 54 Copepoda 85 135 106 10 69 5 436 Diptera Ceratopogonidae Probezzia (I) 1 1 3 1 6 Diptera Ceratopogonidae (ad) 1 1 Diptera Ceratopogonidae (p) 1 Diptera Chironomiidae Chironomiinae (1) 36 13 35 25 41 1 240 Diptera Chironomiidae Chironomiinae (p) 23 3 2 37 3 93 Diptera Chironomiidae Orthocladiinae (1) 35 27 18 33 30 6 300 Diptera Chironomiidae Orthocladiinae (p) 28 16 5 6 4 4 112 Diptera Chironomiidae Tanypodiinae (1) 23 26 23 3 37 3 212 Diptera Chironomiidae Tanypodiinae (p) 81 3 1 226 2 441 Diptera Chironomiidae (ad) 3 2 8 8 34 Diptera Chironomiidae (p) 1 1 Diptera Culiciidae (ad) 12 1 5 1 1 26 Diptera Dixidae Dixa (1) 10 21 2 5 75 Diptera Dixidae Dixella (1) 1 1 Diptera Dixidae Meringodixa (1) 2 15 6 1 3 31 Diptera Empididae Chelifera (1) 2 3 2 7 Diptera Empididae Oreogeton (1) 1 1 Diptera Phoridae (1) 1 1 Diptera Phryganeidae (1) 4 4 Diptera Psychodidae Pericoma/ Telmatoscopus (1) 7 1 5 4 111 Diptera Psychodidae (1) 2 Diptera Psychodidae (pup) 4 1 6 Diptera Simuliidae Prosimulium (1) 4 5 Diptera Simuliidae Simulium (1) 61 2 98 51 65 357 Diptera Simuliidae Simulium (1) 4 12 22 Diptera Simuliidae Simulium (p) 1 1 3 13 Diptera Simuliidae (ad) 22 10 39 5 89 Diptera Simuliidae (1) 7 2 1 14 Diptera Simuliidae (p) 2 2 Diptera Syrphidae Stratiomys (1) 2 2 Diptera Tipulidae Dicranota (1) 4 1 1 12 Diptera Tipulidae Phalacrocera (1) 1 1 Diptera Tipulidae Tipula (1) 2 1 12 Diptera Tipulidae Triogma (1) 3 2 5 Diptera Tipulidae (ad) 3 5 1 16 Diptera Tipulidae (1) 1 1 1 5 Diptera (ad) 32 11 12 5 1 17 80 Diptera (ad) 567 36 148 55 556 79 2266 Diptera (ad) 9 2 8 6 9 52 Diptera (1) 4 2 5 1 2 18 Diptera (p) 1 1 Ephemeroptera Ameletidae Ameletus (n) 1 46 15 61 41 164 Ephemeroptera Ameletidae Parameletus (n) 26 26 Ephemeroptera Baetidae Baetis (n) 3 3 104 34 71 55 504 Ephemeroptera Caenidae Caenis (n) 37 41 Ephemeroptera Caenidae (n) 3 3 16 Ephemeroptera Ephemerellidae Drunella (n) 1 1 Ephemeroptera Ephemerellidae Drunella grandis (n) 12 4 44 3 29 30 122 Ephemeroptera Ephemerellidae Ephemerella (n) 2 1 1 2 10 Ephemeroptera Ephemerellidae Serratella (n) 18 49 5 75 Ephemeroptera Ephemerellidae (n) 3 1 1 1 7 Ephemeroptera Heptageniidae Epeorus (n) 8 2 12 Ephemeroptera Heptageniidae Cinygma (n) 2 1 9 13 Ephemeroptera Heptageniidae Cinygmula (n) 13 2 34 5 54 Ephemeroptera Heptageniidae (n) 13 13 Ephemeroptera Leptophlebiidae Paraleptophlebia (n) 2 15 66 30 11 33 198 Ephemeroptera Leptophlebiidae (n) 5 1 9 1 2 3 21 Ephemeroptera (ad) 3 11 19 11 57 Ephemeroptera (ny) 1 3 1 4 1 13 Fish 1 1 5 Gastropoda 149 24 22 8 1 2 221 Hemiptera Corixidae Hesperocorixa (ad) 1 4 5 Hemiptera Corixidae (imm) 16 6 22 Hemiptera Gerridae Limnophorus (ad) 1 1 Hemiptera Gerridae (imm) 1 1 Hemiptera (ad) 2 4 Hemiptera (imm) 5 1 2 9 Hemiptera (imm) 1 1 1 3 Hirudinea 1 1 2 Homoptera Aphididae (ad) 1 3 Homoptera (ad) 17 2 11 2 3 2 38 Homoptera (imm) 5 5 Hymenoptera Formicidae (ad) 12 1 17 Hymenoptera (ad) 12 11 10 6 23 21 174 Lepidoptera Cossidae Prionoxystus (1) 1 2 Lepidoptera Geometridae (1) 1 4 1 9 Lepidoptera (ad) 1 Lepidoptera (1) 1 1 1 2 15 Megaloptera Sialidae Sialis (n) 1 3 Megaloptera (n) 1 1 Nematoda 1 1 1 1 11 Neuroptera (ad) 1 Odonata (imm) 4 2 1 8 Oligochaeta 38 30 40 17 5 2 165 Ostracoda 268 445 21 11 265 230 1475 Pelecypoda 43 4 1 1 2 60 Plecoptera Capniidae (n) 1 1 3 1 9 Plecoptera Chloroperlidae Sweltsa (n) 6 2 27 13 83 Plecoptera Leuctridae Despaxia (n) 1 9 1 11 Plecoptera Leuctridae (n) 1 1 1 3 Plecoptera Nemouridae Amphinemoura (n) 25 25 Plecoptera Nemouridae Zapada (n) 8 3 1 20 30 15 190 Plecoptera Nemouridae (n) 1 1 Plecoptera Perlidae Hesperoperla (n) 6 Plecoptera Perlidae (n) 2 17 Plecoptera Perlodidae (n) 3 9 34 9 9 4 75 Plecoptera (ad) 3 2 7 2 6 3 34 Tadpoles 2 85 Thysanoptera Thripidae (ad) 4 33 3 9 9 58 Trichoptera Brachycentridae Micrasema (1) 3 11 7 21 Trichoptera Glossosomatidae Glossosoma (1) 1 2 7 10 Trichoptera Hydropsychidae Arctopsyche (1) 1 1 Trichoptera Hydropsychidae Hydropsyche (1) 2 1 20 Trichoptera Hydropsychidae Parapsyche (1) 1 1 Trichoptera Hydropsychidae (1) 1 2 2 8 Trichoptera Hydroptilidae (1) 2 26 1 6 35 Trichoptera Lepidostomatidae Lepidostoma (1) 2 18 Trichoptera Limnephilidae Cryptochia (1) 2 2 Trichoptera Limnephilidae Dicosmoecus (1) 2 1 3 6 Trichoptera Limnephilidae (1) 20 23 2 22 39 134 Trichoptera Rhyacophilidae Rhyacophilia (1) 3 5 6 10 28 Trichoptera (ad) 2 1 4 4 24 Trichoptera (1) 6 26 3 2 131 Trichoptera (p) 1 2 Unknown terrestrial 1 1 2 1 10 Grand Total 2412 1305 1399 481 2248 961 12931 Figure 1.1 Map of the study area showing the locations of the three streams, the upstream and downstream sites on those streams, and their associated cutblocks (in the case of the two treatment streams) (from Mellina 2002). Figure 1.2. Mean (±1 standard error) biomass (a) and abundance (b) of benthic invertebrates before logging (1997, black) and after logging (2001, grey) in the treatment streams (118/16 and 118/48) and stream Hip Creek. Asterisks indicate significant differences. 30 Figure 1.3. Canonical variates scores for benthic invertebrate community composition by biomass (a), and abundance (b). Polygons denote maximum range of scores. Percentages indicate amount of total variation explained by specific canonical variates axes. Streams are indicated by differing symbols: stream 118/16 (•), Stream 118/48 (A), and Hip Creek (•). For biomass (a), group centroids all differ from one another. For abundance (b), stream 118/16 differs from 118/48 and Hip, the latter do not differ. a) b) to 2 5 s •§ a <D O CO 8- & §• o o m - c S O 0_ h- _J Z < o 2 — a> w c E _ co x: t o a. >• < LJ Z | 4-3 * \ / n = 16 \ v o ' n = i o . : - 2 -•3-.A-I ^ 2 4 / ( Acarina Coleoptera Larvae Plecoptera Nymphs CAM (82%) Diptera Larvae Pelecypoda Other Aquatic (0 ~ & 3 3 3 _ j j s « (D <D S O" _ _ 9-< O O ° * -_ _ •£ <B o o -c S O O h O CM z < O CD a) co J_ E ro o CD C O x: o cn UJ O O y y / L 2 \ ^ I n = 16 \ \ \ / 5 -A J -3 -2 \ -1 ( \ n = 12 1 a_ ' ~~ - - V I 1 2 .3. - - • 4 ! . . - * ;» »-*-"-.. n = 10 ,-** Oligochaeta Trichoptera CAN1 (81%)" Coleoptera Larvae Collembola Megaloptera Plecoptera 31 Figure 1.4. Canonical variates scores for functional feeding guild structure between streams using (a) biomass (• 118/16a A118/48 a b • Creekb) and (b) abundance (•I 18/16a A118/48 b • Hip Creekb). Superscript letters indicate significant differences between group centroids. Polygons denote the maximum range of scores. Percentages indicate the amount of total variation explained by specific canonical axes. a) I I O o Collector-Gatherers Collector-Filterers Shredders - - ^ " ' CAN1 (77%)-Scrapers Parasites b) CD M! <5 S-o CO O 0. < O o _ o o . G O CO 2r5-/ V * * < *•» » \ \ 1 / -3 -2 -1 V ( 1 \ -0.5 \ -1.5 I •—— -2T5-- - - - • \ ' 1 1 ^ 3 4 \ / / / _ 4. Shredders Parasites/Piercers CAN1 (88%) ^Predator-Engulfers "Scrapers Figure 1.5. Mean (±1 standard error) drift invertebrate biomass per cubic meter of water by stream, site, and month. Asterisks indicate significant differences between class levels. 33 Figure 1.6. Mean (±1 standard error) total drift invertebrate abundance in the logged streams 118/16,118/48, and unlogged control Hip Creek in July (black) and August (grey) 2001. Asterisks indicate significant differences among months. 20 n co~ 18 -' E * 16 -118/16 118/48 Hip 34 Figure 1.7. Canonical variates scores for biomass of drift taxa by (a) site within stream (•118/16 DS a #118/16 US b A118/48 DS A118/48 US "Hip Creek DSC • Hip Creek US d), and (b) month within stream. (•118/16 DS a •118/16 US b A118/48DS A118/48 US • Hip Creek DS° • Hip Creek US d). Superscript letters indicate significant differences between sites. Polygons denote the maximum range of scores. Percentages indicate the amount of total variation explained by specific canonical axes. a) b) co .JS § • • § « § E s e t CO © «D _ U 1 Q U O < S o .9- ta Q CD ™ • - • o ,'*4-m 3 • a • . 1 --* ." „ , M I a— < Ephemeroptera Plecoptera CAN1 (41%)- Diptera Pupae Other Aquatic Ostracoda Araneae Diptera Adults Gastropoda .CAM (40%) Collembola Other Terrestrial Hymenoptera Adults 35 a) Figure 1.8. Canonical variates scores for abundance of invertebrates within drift taxa between sites within streams between July (a), and August (b). Percentages indicate the amount of total variation explained by specific canonical axes. Polygons denote the maximum range of scores. Sites within streams are indicated by differing symbols: •118/16 DS •118/16 US A118/48 DS A118/48 US "Hip Creek DS "Hip Creek US. b) s §31 I l l s ! . < O O O f z < s I a> _j i2 o § J E &S. | 8-/ / / / / J A - -F 'L a , ~ -*m ..• A i-'''!M' 2 -10 -8 •€ -4 -2 f) f l i' _4-2 4 6 s / / s i 1 2 , £ ca <5 s s a 8" i o c < O Q O C z 8-6 -4 -2 -2 -10 -8 -6 -4 -2 'R 4_ i j / ' 2 4 6 1 I P * D i p t e r a A d u l t s H y m e n o p t e r a A d u l t s -CAN1 (57%)-C o l e o p t e r a L a r v a e D i p t e r a L a r v a e O t h e r T e r r e s t r i a l D i p t e r a A d u l t s H y m e n o p t e r a A d u l t s -CAN1 (57%)- C o i e o p t e r a L a r v a e D i p t e r a L a r v a e O t h e r T e r r e s t r i a l 36 Figure 1.9. Canonical variates scores for drift composition (by biomass) by functional feeding guild. Sites within streams are indicated by differing symbols: •118/16 DS a 4118/16 US b A118/48DS A118/48US "Hip Creek DS ^Hip Creek US. Superscript letters indicate significant differences between sites. Percentages indicate the amount of total variation explained by specific canonical axes. Polygons denote the maximum range of scores. CL CO Scrapers Shredders/Piercers CAN1 (59%) Collector-Gatherers Collector-Filterers Terrestrial 37 Figure 1.10. Canonical variates scores for drift functional feeding guild abundance between sites within streams between July (a) and August (b). Polygons denote the maximum range of scores. Percentages indicate the amount of total variation explained by specific canonical axes. Sites within streams are indicated by differing symbols: •118/16 DS •118/16 US A118/48 DS A118/48 US "Hip Creek DS • Hip Creek US. a) e_. 5 • 4 • 3 • 2 --—' ! i -3 -2 V » J 1 " » ' - . '< -V--2 -3 --4 • _ - C. . . K. *Nl 2 3 4 5 1 C o l l e c t o r - F i l t e r e r s P r e d a t o r s T e r r e s t r i a l -CAN1 (47%) ^ C o l l e c t o r - G a t h e r e r s ' P a r a s i t e s S c r a p e r s b) 6-5 -4 3 2 ja A-' \ . . . . . ; J » -A \ >"•* r " \ / -4 • e 2 " i co b_ 9 "8 O CO C o l l e c t o r - F i l t e r e r s P r e d a t o r s Te r res t r i a l -CAN1 (47%) C o l l e c t o r - G a t h e r e r s P a r a s i t e s S c r a p e r s 38 Chapter 2: The effects of forest harvesting on rainbow trout (Oncorhynchus mykiss) diet in sub-boreal streams of British Columbia INTRODUCTION Streamside timber harvesting has varied effects on salmonids (Oncorhynchus spp.) and their habitats. Deleterious effects include increased sediment and debris loads (Burns 1972; Platts et al. 1989), increased nutrient input from overland runoff (Hansmann and Phinney 1973; Murphy and Hall 1981; Noel et al. 1986; Gregory et al. 1987), increased solar radiation leading to increased temperature (Platts and Nelson 1989; Hetrick et al. 1998b), and decreased habitat stability and complexity (Chamberlin et al. 1991). However, the increased nutrients and solar radiation can cause an increase in stream primary productivity (Murphy and Hall 1981; Murphy et al. 1981; Shortreed and Stockner 1983; Murphy et al. 1986; Beschta et al. 1987; Anderson 1992; Stone and Wallace 1998), which can lead to increased benthic invertebrate density (Hawkins et al. 1982; Stone and Wallace 1998), biomass, and diversity (Newbold et al. 1980; Murphy et al. 1981; Stone and Wallace 1998). The effects of increased invertebrate production may be mirrored in fish growth and production (Murphy and Hall 1981; Hetrick et al. 1998a), provided that stream temperatures remain below the lethal limits of the species. In small streams, drifting benthic and terrestrial invertebrates make up a large portion of the diets of stream-resident salmonids (Bryan and Larkin 1972; Bisson 1978; Allan 1981; Glova 1984; Angradi and Griffith 1990); therefore, it is important to understand how invertebrates and fish respond to environmental disturbances. A number of factors may influence prey choice in salmonids including prey density (Jenkins et al. 1970; Ware 1972), prey size (Ware 1972; Allan 1981; Angradi and Griffith 1990), and predator size (Bisson 1972). Some studies have reported that terrestrially-derived invertebrates comprise up to 30% of the invertebrates in fish stomachs and up to 50% of the biomass (Jenkins et al. 1970; Wipfli 1997). Other studies suggest that the 39 terrestrial component of trout diet is small (Angradi and Griffith 1990; Dedual and Collier 1995). Furthermore, additional investigations have shown that some trout resort to benthic feeding when drift is inaccessible due to turbidity (Tippets and Moyle 1978) or low light (Angradi and Griffith 1990). The single most important factor that seems to regulate prey selection by trout is prey size (Bisson 1978; Allan 1981; Angradi and Griffith 1990). Allan (1981) found that larger taxa were consistently over-represented in the diets of trout, and Bisson (1978) reported that trout fed on large, but relatively rare invertebrates, missing out on the opportunity to exploit small Collembola, which dominated the drift. Regardless of the details, most of these studies reveal a close association between stream-dwelling trout and the benthic invertebrate community. In all of these studies, aquatic invertebrates make up the bulk of trout diet, thus it is important to know how this food source is affected by forest harvesting. Mellina (2002) in a survey of BC central interior streams, found that trout in warmer streams emerged earlier and experienced enhanced growth relative to those in cooler streams, regardless of whether logging was present. In his study, enhanced water temperatures that brought the streams into the optimal temperature range for rainbow trout growth, coupled with increased incoming solar energy and primary production were thought to be the main mechanisms behind the trend in fish size. In the previous chapter, my results indicated that benthic and drift invertebrate abundance, biomass, and community composition did not differ following the application of the logging treatment. Thus, increasing solar radiation leading to enhanced primary and secondary production are likely not the mechanisms behind the enhanced trout growth observed by Mellina. A number of researchers have reported that light levels significantly affect prey detection rates by salmonids in streams (Wilzbach et al. 1986; Young et al. 1997). While the removal of streamside timber has not affected the abundance or biomass of prey items within the streams, it is possible that the increased light associated with logging has increased the ability of resident trout to detect prey items. 40 I investigated the relationship between forest canopy cover, invertebrate drift, and rainbow trout diet. The main objective was to explain the increased condition of the rainbow trout in the treatment streams found by Mellina (2002). I compared fish gut contents from logged and unlogged streams to examine the abundance of prey items and to characterise rainbow trout diet in these streams. I also compared the diets of the fish to the contents of the drift to test for prey selectivity. This allowed the examination of the following hypotheses: a) that rainbow trout in the logged streams experience greater prey detection rates and are able to feed on a greater number of invertebrates than those in the unlogged control streams; and b) that trout feed selectively on larger size classes of invertebrates. MATERIALS AND METHODS I examined three small, lake-headed streams (118/16, 118/48, and Hip Creek, Figure 1.1) located within the sub-boreal spruce biogeoclimatic zone of north-central BC (Farley 1979). This region is dominated by glaciolacustrine and sandy glaciofluvial soils (MacDonald et al. 1992, Larkin et al. 1998). Annual precipitation averages 50cm, falling primarily as snow in winter, with seasonal peaks in late spring and early autumn MacDonald et al. 1992). The hydrograph is dominated by spring meltwater runoff (Cheong et al. 1995); however, a smaller peak occurrs in autumn due to increased rainfall (Beschta et al. 1987). The streams were matched as closely as possible with respect to their physical characteristics (see details in Mellina 2002). Each had a bankfull width of less than 2 m, a gradient lower than 4%, and was primarily dominated by pool-riffle morphology (Table 1.1). All of the streams were headed by small, shallow lakes, although Hip Creek had a wetland complex situated between the lake and the study site. On each stream, a study section 400 to 600 m in length was delineated by upstream (US) and downstream (DS) sampling sites. No tributaries occurred within the study sections. Fish species composition in all streams was dominated by rainbow 41 trout (Oncorhynchus mykiss), although small numbers of suckers (Catostomus spp.), lake chub (Couesius plumbeus), prickly sculpins (Cottus asper), and burbot (Lota lota) were present (Mellina 2002). Each stream was studied one year before logging (1997) and three years after riparian logging (2001). Logging was conducted by Canadian Forest Products in February 1998. The riparian portions of upland around streams 118/16 and 118/48 (hereafter termed the "treatment streams") were clear-cut, and all merchantable timber (>15 cm diameter at breast height (DBH) for lodgepole pine (Pinus contorta) and >20 cm DBH for spruce (Picea spp.) and sub-alpine fir (Abies lasiocarpa)) was removed. A 5 m buffer from which heavy machinery was excluded was maintained around treatment streams to preserve the soil quality and bank stability. Fish sampling sites were located within the upstream and downstream boundaries of the clear-cuts on the treatment streams. Clearcut size was 40 ha for 118/16 and 36 ha for 118/48. Hip Creek served as a control stream and was not subjected to any forest harvesting treatment. To obtain estimates of growth and condition of rainbow trout in the study streams, minnow traps were used to conduct a mark-recapture study. Minnow traps were baited with canned cat food wrapped in cheesecloth, and placed in twenty locations throughout the study area. The traps were placed in the streams at marked habitat units just before the start of drift sampling (see chapter 1), and removed one day after drift sampling was completed for, a total sampling time of approximately 48 hours. Traps were placed in the same locations in both months. Once the trout were removed from the traps, they were anaesthetised with tricaine methanesulfonate (MS-222), measured, weighed, and implanted with a passive integrated transponder (PIT) tag. Trout were then replaced in the habitat where they were caught. Following the methods outlined in Elliott and Persson (1978), each stream was electrofished for fish four times over the 24-hour period (dawn, noon, afternoon, and dusk intervals). We 42 included dusk and dawn sampling intervals as this is often suggested to be the time of peak feeding activity due to increased invertebrate drift. During each sampling interval, the selected pools we electroshocked were completely emptied of fish. Five trout from each sampling interval were first anaesthetised with MS-222, and then measured and weighed, resulting in a total of 20 trout per 24-hour period per stream. Fish were anaesthetised, and water was pumped into their stomach via the mouth to flush out the stomach contents (Culp et al. 1988, Meehan and Miller 1978). Due to difficulties encountered while handling small fish, fish less than 5 cm in length were sacrificed and their gut contents were removed in the lab. All samples were preserved in 10% formalin, counted, and identified to the lowest possible taxonomic level (usually family) in the lab. For each fish, the length of each invertebrate found in its stomach, as well as the total biomass of the entire stomach content, was estimated using the ZooBiomass v1.31 (Hopcroft 1995) digitising program. This software uses published length-weight relationships (Smock 1980; Johnston and Cunjak 1999) to estimate the dry weight of each invertebrate. Because many of the invertebrates in the stomachs were partially digested, this method likely provides a better estimate of stomach content biomass than direct weighing of preserved specimens. For prey items that were digested to a point where they could not be measured, head capsules were counted, and the mean dry mass estimated from measured organisms in the same sample were applied. I had originally intended to use length and weight data from individual PIT-tagged trout recaptured in August to generate estimates of summertime growth. However, growth rates could not be accurately estimated using this method because sample sizes were extremely low (only one fish tagged in July was recaptured during the August sampling period). Therefore, length-mass relationships were used as a relative measure of fish condition and a surrogate for growth of captured fish (Boss and Richardson 2002). Analysis of variance (ANOVA) was performed to compare the linear regression relationship of fish mass among streams, with fish length as the covariate. Furthermore, because young-of-the-year (YOY) trout had not yet 43 emerged in control stream HIP during the July sampling period, all analyses were restricted to juvenile trout (defined as older than age-0). A detailed census of the trout populations in the case-study streams was not conducted during the sampling season; therefore, I assumed that the population structures in the study streams did not differ substantially between 1999 (the last year for which census data was available) and 2001. Consequently, size limits for YOY trout were obtained from length-frequency distributions constructed using data collected in 1999 by Mellina (2002). ANOVAs were used to assess differences in the total number and biomass of invertebrates found in trout stomachs among streams and among months (July and August). The mean dry mass of individual invertebrates was also examined to test for diet selectivity among rainbow trout. Estimates of individual dry mass of invertebrates in the stomach contents were obtained from the ZooBiomass digitising program. To determine if logging affects prey availability or quality, individual prey item size was compared to the average individual dry mass of drifting invertebrates, obtained by dividing the total biomass invertebrates in each drift sample from Chapter 1, by the abundance of invertebrates counted. ANOVAs were employed to test for differences in the average size of invertebrates among streams, sites (upstream versus downstream), and sample types (drift versus prey). This method was also employed to test for differences in the average size of the three primary prey items (defined as those taxa making up the greatest biomass and abundance within the gut contents). To gauge the effects of the forest harvesting treatment on these preferential prey items, analyses of variance were also employed to test for differences in the total biomass and abundance of the primary prey items in the drift among streams, sites, and months. Bonferroni t-tests were used as an a-posteriori comparison of means for each category. 44 Multivariate analysis of variance (MANOVA) was employed to assess differences in the biomass and abundance of prey items in the composition of fish gut contents among streams, sites, and months. If MANOVA identified that site, stream, month, or any interactions among these classes were significant, canonical variates analysis (CVA) was performed to assess the differences between multivariate centroids, and the scores were plotted to visualize the results. For all analyses, a significant interaction effect would indicate that a change in one of our invertebrate variables (i.e. prey abundance or biomass) from upstream to downstream sites differed among streams, which indicates that the logging treatment may have had an effect. Data were log10(x,+ 1) transformed to meet the assumptions of normality required for parametric tests. For all analyses, results were considered significant if the probability of a false significant result was less than 0.05 (i.e. 95% confidence). SAS software (SAS Institute Inc., 1988) was employed for all analyses. RESULTS The interaction of fish length by stream did not differ among streams (GLM, P = 0.31). Thus, weight at a given length (indicated by the value of the y-intercept) of trout could be compared between streams. The results indicate that fish from Hip Creek had significantly higher condition than fish from stream 118/16 (ANOVA, P = 0.02; Figure 2.1). The condition of fish from stream 118/48 did not differ significantly from either stream due to the relatively lower R2-value. The total weight of the gut contents of the fish ranged from 0 to 378 mg (dry) and differed significantly by month and by stream (ANOVA, P < 0.01 for both classes). Hip Creek fish had higher stomach content biomass than fish from the two treatment streams. Stomach content 45 biomass was higher in July than August across all streams. The numerical abundance of invertebrates flushed from the fish differed significantly among streams, and among sites (P < 0.01 for both classes). Fish from Hip Creek ate a greater number of invertebrates than those from 118/16 and 118/48 (Figure 2.2). The composition of the gut contents by biomass differed between streams and between months within streams (MANOVA; P < 0.01). Canonical variates analysis indicated that fish from Hip Creek ate different types of invertebrates than fish from streams 118/16 and 118/48 (CVA; P < 0.01; Figure 2.3a). The first canonical axis explained 63% of the variation, while the second canonical axis explained 18%. Fish from Hip Creek scored high on the first canonical axis, indicating that they ate a greater biomass of Plecoptera and terrestrial invertebrate taxa than those in 118/48 and 118/16. The stomach contents of fish from 118/48 differed from those in 118/16 mainly along the second canonical axis, revealing that fish in stream 118/48 ate a higher biomass of Coleoptera adults, whereas fish from stream 118/16 ate a higher biomass of Araneae, Coleoptera and Trichoptera larvae, and Diptera pupae. The diets of fish from Hip Creek and stream 118/16 also differed between months. In July, fish from Hip Creek ate a larger biomass of Plecoptera nymphs and rare terrestrial invertebrates than they did in August. The diets of fish from stream 118/16 shifted from a high biomass of mites, Coleoptera and Trichoptera larvae, and Diptera Pupae in July, to a diet richer in Coleoptera adults in August (Figure 2.3a). The composition of the gut contents calculated by numerical abundance also differed between streams and between months within streams (MANOVA; P < 0.01). Canonical variates analysis revealed that diets differed mainly along the first canonical axis, which represented 59% of the variation (Figure 2.3b). High scores on the first axis represented a high abundance of Ephemeroptera nymphs, Diptera larvae, Coleoptera adults, and terrestrial invertebrates while 46 low scores represented a higher abundance of Diptera adults and pupae, Coleoptera larvae, and Plecoptera nymphs in the diet. Hip Creek scored highest on this axis, followed by stream 118/48 and stream 118/16. Diets of fish from stream 118/48 scored slightly lower on the second canonical axis (representing 22% of the variation) than the other two streams. This represents a slightly greater number of Diptera larvae and rare aquatic taxa in the diets of these fish. Only the diets of fish from stream 118/48 differed between July and August. Fish from this stream ate more Diptera adults and pupae, Coleoptera larvae, and rare aquatic taxa in July, and more Ephemeroptera nymphs, Coleoptera adults, Diptera larvae, and terrestrial insects in August. The mean dry mass of individual drifting invertebrates (i.e. those available for fish consumption) ranged from 1.5 to 2.0 mg, and was higher in stream 118/48 than in Hip Creek and stream 118/16 (ANOVA, P = 0.01, Figure 2.4). The mean dry mass of individual invertebrates in the stomach contents of fish was higher in Hip Creek than in stream 118/16 (P = 0.02). The average dry weight of individual invertebrates in the diets of fish from stream 118/48 did not differ from the other two streams. The interaction between stream and sample type had an effect on the mean individual weight of invertebrates obtained in this study. Fish from Hip Creek ate larger invertebrates than were generally available in the drift, whereas fish from the logged streams ate invertebrates that were smaller than those available in the drift (ANOVA, P = 0.03). Ephemeroptera and Diptera were the dominant prey taxa in the diets of rainbow trout from these streams. Ephemeroptera (mayfly) nymphs represented the highest biomass of a single taxon in the stomach contents of fish from all streams (average 1.22 to 5.21 mg/fish); except for stream 118/16 in August, when adult Dipterans (true flies) made up the highest biomass (average 0.50 mg/fish). Adult and larval Dipterans also made up large proportions of the stomach content biomass (average 0.42 to 1.38 mg/fish and 0.24 to 4.1 mg/fish respectively). 47 Diptera larvae and adults were the most numerically abundant organisms in the diets of trout in most of the streams, ranging from an average abundance of 1.2 to 18.2 organisms per fish. Ephemeroptera nymphs were also abundant in the diets of fish from Hip Creek, and from the upstream site of stream 118/48 (an average of 2 to 6.2 organisms per fish). The mean individual size of Ephemeroptera nymphs differed among sample types (ANOVA, P < 0.01), but not among streams (P = 0.59) or sites (P = 0.09). The mean dry mass of individual mayflies obtained from the gut contents was heavier than the dry mass of mayflies obtained from the drift samples (Figure 2.5). Similarly, the average size of larval Diptera was also influenced by sample type (P < 0.01), with drifting individuals weighing less than those captured by fish. The mean dry mass of Diptera adults did not differ among any of the variables (P > 0.05 for all variables). There were no interactions between sites and streams that would indicate that any of the primary prey items were affected by the logging treatment (ANOVA, P > 0.05 for all primary prey taxa). Biomass and abundance of Ephemeroptera in the gut contents were affected by stream (P = 0.02) and month (P = 0.02), with the guts of fish from Hip Creek containing the highest number and biomass of mayflies, followed by stream 118/48, and stream 118/16 (Figure 2.6). There was also a weak but significant effect of month within stream on the biomass of Ephemeroptera within the gut contents (P = 0.03). The total biomass of mayflies in the gut contents was higher in July than August in stream 118/16 and Hip Creek, but not in stream 118/48 (Figure 2.7). The total abundance of Ephemeroptera in the gut contents was highest in Hip Creek, followed by streams 118/48 and 118/16 (ANOVA, P < 0.01), and higher in July than August (P = 0.02). Within the drift, total Ephemeroptera biomass was not affected by stream (ANOVA, P = 0.69), site (P = 0.82), month (P = 0.32), or any interaction of those variables (ANOVA, P>0.90 for all interactions). Ephemeroptera abundance in the drift was affected only 48 by month, and was higher in July (1.30 individuals/m3) than in August (1.15 individuals/m3; P = 0.04). The total biomass of Diptera adults in the gut contents did not differ significantly among streams (ANOVA, P = 0.20), sites (P = 0.11), months (P = 0.88), or any interaction of these variables (ANOVA, P > 0.10 for all interactions); however, their abundance differed among streams, with fish in Hip Creek eating a greater number of Diptera adults than fish in stream 118/48 (ANOVA, P = 0.01). Similarly, Diptera larvae did not vary among streams (ANOVA, P = 0.47) or months (P = 0.18) in their contribution to stomach content biomass; however, fish in all streams ate a higher biomass of Diptera larvae at the upstream sites than at the downstream sites (P = 0.01). The abundance of Diptera larvae in the diets offish differed among streams (P < 0.01), sites (P < 0.01), and between sites within streams (P < 0.01). In general, fish ate greater numbers of Diptera larvae at the upstream site than the downstream site, except for in stream 118/48 where a greater number were eaten at the downstream site than the upstream site. In the drift, there was a difference between months within stream on the biomass and abundance of Diptera adults and larvae. During July sampling, abundance and biomass of Diptera adults and larvae were higher in stream 118/16 than in stream 118/48 and Hip Creek. In August, there was no significant difference in the biomass or abundance of Diptera adults and larvae among streams (ANOVA, P>0.05 for all comparisons). DISCUSSION Rainbow trout in the control stream were heavier across all lengths, and ate a greater number and biomass of invertebrates than trout from the logged streams, suggesting that the reduction in canopy cover following logging did not increase prey detection rates as expected. This contrasts with conclusions from Mellina's (2002) comparative survey of sub-boreal B.C. 49 streams. In his study, Mellina (2002) found that rainbow trout from streams that had been subjected to clearcutting up to thirty years prior to the study had higher condition relative to fish in unlogged streams. Murphy and Hall (1981) found a greater biomass, density, and species richness of predators (including trout) in coastal streams exposed to direct sunlight (i.e. within 5 to 17 years after logging). In both of these studies, enhanced water temperatures that brought the streams into the optimal temperature range for rainbow trout growth, coupled with increased incoming solar energy and primary production were thought to be the main mechanisms behind the trend in fish size. In addition, Thedinga et al. (1989) found that coho fry experienced increased size and condition in logged streams; however, age-1 parr did not differ in size or condition between logged and unlogged streams. This implied that the fry experienced increased growth because increased water temperature and food abundance resulted in their earlier emergence in the summer in the logged streams, giving them more time to grow before the onset of winter. By the time one year had passed, the benefits of this early emergence had been lost, presumably due to the negative effects of increased stream temperatures on growth rates (Thedinga et al. 1989). I suggest that the lower condition of fish in my logged streams was a result of the lower number and biomass of invertebrates ingested, and that this phenomenon may have stemmed from a combination of the lower drift invertebrate abundance (chapter 1), and high water temperatures within these streams. In addition, higher densities of trout in the logged streams may have played a role in decreasing the overall condition of fish. While density-dependent growth is more common in lakes than in streams (Post et al. 1999), some degree of density-dependent growth may occur, since fish must defend their territories more avidly, and therefore may acquire less food or burn energy more quickly (Bohlin et al. 1994; Jenkins et al. 1999; Vollestad et al. 2002). In addition, Slaney and Northcote (1974) found that territorial behaviour decreased and trout density remained high when rainbow trout were introduced to stream channels with high prey density. Conversely, when prey levels were low, territorial behaviour increased and 50 trout density decreased. In this study, invertebrate drift abundance and biomass was lowest, and trout density was highest in the two treatment streams. Mellina (2002) found that streams 118/48 and 118/16 had a significantly higher density of fish than Hip Creek. In addition, these fish tended to aggregate in suitable habitats, whereas fish in Hip Creek were distributed randomly. During sampling, three to five fish were regularly captured in each habitat unit of the logged streams, whereas in the control stream, only one to two fish per habitat unit were caught. Thus, intraspecific competition for food in streams 118/48 and 118/16 resulting in increased aggression, increased activity, and decreased prey detection may be contributing to the lower condition of the fish observed in these streams. Temperature increases following logging are well-documented both in coastal (e.g. Johnson and Jones 2000), and interior (e.g. Macdonald et al. 2003) regions of western North America. In general, warmer stream temperatures increase the metabolism of fish, prompting them to acquire more food, which can then be converted into growth. This holds true up to a maximum temperature (defined by species and population), at which point the maintenance cost associated with the high metabolic rate exceeds the amount of food that can be acquired, leading to decreased growth and eventually mortality (Brett and Groves 1979). Rainbow trout have been shown to be tolerant of a wide range of temperatures in laboratory experiments (0-29°C; Reiser and Bjornn 1971, Hokanson et al. 1977, Lee and Rinne 1980). Growth rate is generally high between 10 and 20°C (Railsback and Rose 1999), and likely reaches optimal levels between 15.5 and 17.3°C (Hokanson et al. 1977). The average monthly maximum water temperature was highest in stream 118/16 in July (18.7°C), and was approximately 2.3° higher than the average July maximum in stream 118/48 (16.4°) and 4.9° higher than the July maximum in Hip Creek (13.8°; Mellina unpublished data). August data showed similar, but less extreme trends. In addition, stream 118/16 is the only stream where temperatures exceed the lethal published maximum for salmonids (23-25°C; Scott and Crossman 1973) in the summer (Mellina et al. 2002). From a bioenergetics perspective, these high temperatures may suggest 51 that fish in this stream must expend more energy on maintenance and less on growth. Hokanson et al. (1977) suggested that rainbow trout could survive lethal temperatures for short periods; however, the increased metabolic costs associated with very high water temperatures would result in negative impacts on growth. Thus, even though trout in stream 118/16 spent a long period within the optimal temperature range for growth relative to Hip Creek (Mellina 2002), the periods of extreme temperatures coupled with increased activity due to competition may negate the potential benefits of the time spent within the optimal temperature range. Stream 118/48 had the highest number of hours within the optimal temperature range for rainbow trout growth and did not exceed lethal limits (Mellina 2002); however, intraspecific competition may erase the potential benefits of these preferential temperatures. Finally, although Hip Creek spends the least amount of time within the optimal temperature range, the significantly lower density and the random distribution of fish among habitats tends to proffer a benefit with regards to the amount of food and space available and hence, to growth. While canopy removal was not responsible for the warm temperatures observed in these three streams, the data presented in this study demonstrate that harvesting practices which bring streams into the temperature ranges observed in streams 118/16 and 118/48 may be detrimental to rainbow trout and other salmonid species. Fish in Hip Creek (control) had a greater number and biomass of invertebrates in their stomachs than fish in the logged streams, despite the fact that the logged streams (specifically Stream 118/16) contained a greater abundance and biomass of invertebrates in the drift. Similarly, Angradi and Griffith (1990), and Tippets and Moyle (1978) found that trout stomach fullness was not correlated to invertebrate drift density in their streams in California and Idaho. These authors postulated that because trout are visual predators, they do not feed at night when drift densities are highest, but rather feed primarily on the bottom of the stream during the day. Bisson (1978) also found that rainbow trout shifted to benthic foraging at night. Epibenthic feeding such as this may have contributed to the differences in fish condition since benthic 52 invertebrate abundance and biomass were similar among streams (chapter 1); however, I do not believe that it fully explains the higher relative condition of fish in the control stream since benthic invertebrate abundance and biomass in that stream were not significantly higher than the logged streams. In addition, other authors have reported that trout stomach fullness and contents were positively correlated with invertebrate drift density and composition (Jenkins et al. 1970; Allan 1981; Cada et al. 1987). Hayes et al. (1995) contended that in studies of length-weight regressions among fish, y-intercepts are often biased high. In my study, sample sizes were similar across all streams, and all fish were captured within one week of each other in both months. Also, the month of capture did not significantly affect the slope or intercept of the length-weight relationships within streams, validating the decision to pool fish data from both months. For these reasons, I feel that if bias occurred in the analyses of length-weight relationships among streams, it was likely similar across all streams; therefore, comparisons of weight-at-length among streams should remain valid. It may also be argued that the length of time that trout were left in the minnow traps may have affected the relative condition of the fish caught by restricting their ability to feed for a long period. Fish at higher temperatures consume energy more quickly than those at lower temperatures, and rates of gastric evacuation increase with increasing temperature (Elliott 1976; Elliott and Persson 1978). However, the use of length-weight estimates (Smock 1980; Johnston and Cunjak 1999) to calculate the biomass of all stomach contents eliminates much of the bias caused by measuring the weight of partially digested items. Temperature differences between streams also affect the relative condition of fish caught in minnow traps, since fish in warmer water will be more active, and burn energy more quickly; thus, a prolonged period without food may have an impact on their body size. Simpkins et al. (2003) studied the effects of fasting on the body composition of sedentary and active rainbow trout. Their findings 53 indicated that the activity level of fasting trout did not affect the relative weight (weight for a given length) of the study specimens for the first 102 days of the study. Since the fish caught in this experiment were only left in the minnow traps for 48 hours, differences in body condition due to stream temperature differences (and hence, differences in fish activity levels) are not probable. The composition of the invertebrate drift community is intimately associated with the composition of the diets of resident fish (Cada et al. 1987; Angradi and Griffith 1990; Dedual and Collier 1995); therefore, it is likely that variables which affect invertebrate composition in streams will also cause differences in diet composition. Macroinvertebrate community composition in these three streams is likely affected by the natural temperature differences between the streams and not by increases in solar radiation and nutrients due to the logging treatment (Chapter 1). The response variables investigated in the previous chapter did not reveal post-logging differences in the invertebrate community among streams, nor did the comparison of drift invertebrate communities in the logged and unlogged streams. Analyses of diet composition revealed that temperature-sensitive organisms such as Plecoptera (stoneflies) and Ephemeroptera made up a greater proportion of the biomass and abundance of trout stomach contents in Hip Creek and stream 118/48; whereas, diet composition of trout in the warmest stream (118/16) was comprised of more temperature-tolerant invertebrate orders such as Coleoptera and Diptera. Terrestrial and aquatic dipterans made up a significant proportion of the abundance of invertebrates in the stomach contents of all trout in the treatment streams, while Ephemeropterans made up a significant proportion of the diet biomass. In addition, Ephemeropterans were among the top three taxa numerically in the diets of fish from Hip Creek through both months, and in stream 118/48 in August. Stonefly and mayfly nymphs are notoriously susceptible to high stream temperatures (Ward and Stanford 1982; Ward 1992; Quinn et al. 1994; Merritt and Cummins 1996; Cox and Rutherford 2000), and their lower 54 relative abundance in the diets of fish from the warmer streams may be another indicator of the high stream temperatures found there. In addition to altering drift and benthic invertebrate community composition, high water temperatures may result in changes to larval insect growth rates, which ultimately affect the size at which insects pupate (Sweeney and Vannote 1978; Vannote and Sweeney 1980). When temperatures are warmed or cooled with respect to the insect's thermal optimum, individuals pupate sooner but at a smaller size, thus reducing adult body size and fecundity (Sweeney and Vannote 1978; Vannote and Sweeney 1980). Since trout have been found to feed preferentially on larger prey items in the drift (Allan 1981, Bisson 1978), stream temperatures may affect the size of the largest prey items and negatively influence the diets of resident fish. In support of this theory, trout from Hip Creek ate invertebrates that were consistently larger than those eaten in the treatment streams, resulting in a higher biomass/abundance ratio of invertebrates within the stomach contents. The average size of prey items selected by the trout did not reflect the average size of organisms available in the drift. Fish in Hip Creek consistently ate invertebrates that were larger than the average size of those in the drift, while fish in the logged streams ate invertebrates that were smaller than the average size of those in the drift. The apparent preferential feeding by fish in Hip Creek may stem from behavioural differences among fish in the study streams. Bisson (1978) found that larger trout tended to exploit large adult chironomids on the surface, while smaller trout fed primarily on midwater drift in the water column. It is possible that fish in Hip Creek eat larger organisms than those in the treatment streams simply because they are larger themselves, and can exploit larger prey items. Alternatively, fish from the two treatment streams may feed more commonly on terrestrial invertebrates such as Diptera adults, which fall into the stream from surrounding vegetation. Streamside logging increases light levels, and 55 may increase terrestrial inputs of invertebrates to small streams (Wipfli 1997). Overall, terrestrially-derived invertebrates (especially Hymenoptera and Diptera adults) made up a high proportion of the drift biomass in stream 118/16, especially in July (Chapter 1), but stream 118/48 did not show such dramatic differences in drift content. Light levels have been shown to positively influence prey detection rates in salmonids (Mazur and Beauchamp 2003; Vogel and Beauchamp 1999). Invertebrates drifting on the surface may not be larger than those drifting in the water column; however, in the absence of a full forest canopy, they are silhouetted against the bright sky, and may therefore be more visible to fish feeding from below the surface (Wilzbach et al. 1986; Young et al. 1997). Trout growth and fish size may be influenced by many factors that are affected by logging, but which could not be measured in this study, such as habitat depth (Harvey and Stewart 2001), habitat complexity (Fausch and Northcote 1992; Roni and Quinn 2001), and overwintering conditions (Cunjak 1996). I contend that in the case of this study, habitat depth and complexity were not affected by the logging treatment due to the short period of time since the logging treatment was applied. Since timber cutting was performed on a deep snowpack, damage to streambanks and in-stream structures was negligible. Furthermore, the streams were small enough, and the flows low enough that loss of functional large woody debris due to flooding was unlikely. In conclusion, this study has revealed that streamside forest harvesting has not directly resulted in impacts to rainbow trout or their diets and food sources. Possible explanations include the unique temperature regimes associated with lake-headed streams, and intraspecific competition among stream salmonids. Trout size and prey visibility may affect the size of invertebrates eaten by trout in streams, and these may be affected by forest harvesting. Unlike other studies which reported increases in food availability in clear-cuts, these data do not suggest that prey 56 availability for stream salmonids changes in the few years immediately following streamside logging. 5 7 Table 2.1. List of invertebrate taxa and their abundances in the diets of rainbow trout from the study streams in July and August 2001. (L = Larvae, Ad = Adults, Pup = Pupae.) Grand Taxa 11816 11848 HIP Total July August July August July August Acarina 7 4 2 3 1 23 Amphipoda 1 1 Arachnida 3 2 1 12 5 33 Baetidae 3 9 26 19 80 Ceratopogonidae 5 3 21 5 65 Chironomidae 3 3 1 2 9 Chironomiinae 1 8 49 29 58 7 163 Chloroperlidae 1 2 17 8 30 Coleoptera 8 4 11 8 87 15 185 Coleoptera L 73 1 1 2 3 3 89 Diptera Ad 167 54 109 35 208 96 953 Diptera L 4 1 17 4 13 61 Diptera Pup 26 24 17 14 2 101 Dytiscidae 1 1 Empididae 1 2 5 1 10 Ephemerellidae 2 5 9 14 8 38 Ephemeroptera 14 3 37 19 23 13 115 Ephemeroptera Ad 5 3 11 2 13 14 68 Formicidae 14 7 28 Glossosomatidae 1 1 2 Hemiptera 1 6 32 1 42 Heptageniidae 3 2 6 2 42 27 83 Homoptera 13 6 14 12 2 63 Hydrophilidae 1 1 1 4 Hydropsychidae 1 2 1 5 Hydroptilidae 1 1 Hymenoptera 8 2 4 7 12 4 69 Lepidoptera 2 1 41 8 77 Lepidostomatidae 1 4 Leptophlebidae 2 2 4 Leuctridae 1 1 6 8 Limnephilidae 1 39 3 43 Nematoda 10 2 13 7 39 Nemouridae 2 1 3 2 5 1 19 Oligochaeta 2 1 1 4 Orthocladiinae 30 30 36 14 128 13 278 Perlidae 1 Plecoptera 4 2 16 7 9 6 49 Rhyacophilidae 1 2 7 11 Sialidae 2 2 Simuliidae 105 5 71 16 204 Tanypodiinae 11 1 23 1 44 58 Tanypodinae 7 2 9 Tipulidae 1 1 2 3 2 17 Trichoptera 2 14 2 11 6 5 42 Trichoptera Ad 2 2 1 3 8 2 22 Grand Total 494 225 357 253 899 316 3199 Figure 2.1. Graph of the length-weight regressions for age-1 fish from 118/16 (O ), 118/48 (A ), and Hip Creek ( • ). Superscript letters indicate significant differences in the y-intercept between streams. Figure 2.2. Mean (±1 standard error) biomass (a) and abundance (b) of invertebrates in the stomach contents of rainbow trout. Letters denote significant differences among streams. 61 Figure 2.3. Canonical variates scores on the biomass (a) and abundance (b) of invertebrate taxa in the stomach contents of rainbow trout. Percentages indicate the amount of total variation explained by specific canonical axes. Polygons denote the maximum range of scores. Months within streams are indicated by differing symbols: 4118/16 July 4118/16 August A118/48 July A118/48 August • Hip Creek July • Hip Creek August. a) b) o> 2 g <f 52 8 « 2 §• 0 ) 0 ) 0 . <o o. S. <o < O H Q . 5 4 Y N / *.-«•-• —-A \ / y lr f \ , -4 • • -5-i 1 1 1 .-• 1 .' '-4^ 1 2 . . . . . e3' 4 -'5 I - , . . . - • • * ' " " CAN1 (63%) -Plecoptera Nymphs Other Terrestrial & z < o 2 2 a> o w ® S i « F ^ a . QJ < U J Z C L U Q O Plecoptera Nymphs Coleoptera Larvae Diptera Adults and Pupae 4 -3 -2 -V \~. ..ss._ <"\ " \ / \ 5 - 4 \ .3 -2 V< - 1 / \ ( i V ^ - v \ i f \ \ ' *.— \ -2* \ \ - 4 -1 .^1 m"3 ^"Y A CAN1 (59%) Ephemeroptera Nymphs Coleoptera Adults Diptera Larvae Other Terrestrial 62 Figure 2.4. Mean (±1 standard error) individual invertebrate weight (mg) in the drift (black) and gut contents of trout (white) in the logged streams 118/16 and 118/48, and in the unlogged control stream Hip Creek. Letters indicate significant differences among streams. 63 Figure 2.5. Mean (±1 standard error) individual dry mass (mg) of the three dominant prey groups: Ephemeroptera (black), Diptera larvae (light grey), and Diptera adults (dark grey) in the drift samples and rainbow trout gut contents. Figure 2.6. Mean (±1 standard error) total biomass and abundance of the total gut contents (dark grey), and of the three primary prey families: Ephemeroptera (medium grey), Diptera adults (light grey), and Diptera larvae (white) in the drift samples and stomachs of rainbow trout from the three study streams. Gut contents Drift contents (0 </> re E o CO o c re •D JO < + -1 BB \ \ i E i _ML~jri-f-, Hip 65 Figure 2.7. Mean (±1 standard error) total biomass of Ephemeroptera in the gut contents of rainbow trout in logged streams 118/16 (dark grey) and 118/48 (light grey), and in control stream Hip (white). Letters indicate differences between months. July August 66 REFERENCES CITED Allan, J.D. 1981. Determinants of diet of brook trout (Salvelinus fontinalis) in a mountain stream. Can. J. Fish. Aquat. Sci. 38: 184-192. 1987. Macroinvertebrate drift in a Rocky Mountain stream. Hydrobiologia 144: 261-268. Allan, J.D., Wipfli, M.S., Caouette, J.P., Prussian, A., and Rodgers, J. 2003. 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