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The invertebrate connection : tracing the value of food subsidies from fishless headwaters to downstream… Reiss, Aya Elaine 2007

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THE INVERTEBRATE CONNECTION: TRACING THE V A L U E OF FOOD SUBSIDIES F R O M FISHLESS HEADWATERS TO D O W N S T R E A M FISH POPULATIONS by A Y A ELAINE REISS B.A., Williams College, 2000 A THESIS SUBMITTED IN PARTIAL F U L F I L L M E N T OF THE REQUIREMENTS FOR THE DEGREE OF M A S T E R OF SCIENCE in The Faculty of Graduate Studies (Forestry) THE UNIVERSITY OF BRITISH C O L U M B I A June 2007 © Aya Elaine Reiss, 2007 I ABSTRACT The role of ecosystem subsidies from Ashless, headwater streams for downstream fish populations was investigated through multiple approaches in the coastal temperate rainforest of British Columbia. Terrestrial invertebrate samples were collected in two streams of differing canopy (broadleaf and coniferous) on a weekly basis to assess seasonal (spring to fall) terrestrial inputs to headwater streams. Thirty-four Ashless headwater streams were also sampled for both drift and terrestrial input of invertebrates over two seasons (spring and summer) and two canopy types (broadleaf and coniferous) to quantify the magnitude of upstream subsidies at different times of year and from headwaters with different land-use histories. Potential fish growth response to headwater subsidies was examined both in a feeding experiment and through bioenergetics modeling. Juvenile cutthroat trout in experimental channels were fed one of seven diets (0-9% of their biomass) to determine if biomass accumulation was proportional to food availability. Bioenergetics modeling was used to perform a sensitivity analysis on parameters influencing fish growth and to validate experimentally-derived growth-response curves. Terrestrial invertebrate inputs increased significantly from May-September and a significant increase in total invertebrate export was observed from invertebrates in the drift from May-July. Quantification of subsidies revealed that streams running through broadleaf canopies exported greater, but not significant, numbers (mean = 1.73 individuals/m3) and biomass (mean = 1.37 mg/m3) of invertebrates than those -adjacent to coniferous ones (mean = 1.49 individuals/m and 1.30 mg/m ). Drift abundance was made up of roughly 83% aquatic and 17% terrestrial invertebrates while biomass was composed of 51% aquatic and 49% terrestrial invertebrates. One-year-old cutthroat trout demonstrated significantly increased relative growth rates with greater food abundance. Sensitivity analysis of model parameters identified that variation in the ratio of terrestrial :aquatic prey caused the greatest change in consumer growth output response at high prey densities. Using the experimentally-derived fish growth response curve, measured headwater invertebrate drift densities, and drift densities reported in local fish-bearing streams, a 12-fold increase in relative growth rate of downstream consumers exposed to headwater resources was predicted. Findings from this study highlight the potentially important contributions of food resources from Ashless headwater streams for downstream consumers. TABLE OF CONTENTS ABSTRACT ii TABLE OF CONTENTS iv LIST OF TABLES vi LIST OF FIGURES viii ACKNOWLEDGMENTS x CO-AUTHORSHIP STATEMENT xi CHAPTER 1: ECOSYSTEM SUBSIDIES AND SERVICES - VALUING RESOURCE FLOW 1 Subsidies are everywhere 1 Consequences of subsidies 3 Ecosystem services provided by subsidies 4 Research Questions and Hypotheses 5 References 8 CHAPTER 2: LINKING HEADWATER SUBSIDIES TO DOWNSTREAM CONSUMERS 11 INTRODUCTION 11 METHODS 18 Study Site 18 Terrestrial Inputs to Headwaters 18 Season and Canopy Data 18 Subsidy export from headwaters 19 Season and Canopy Data 19 Physical Environment 22 Fish responses to variation in headwater drift subsidies 23 Bioenergetics Modeling 25 RESULTS 29 Terrestrial Inputs to Headwaters 29 Subsidy Export from Headwaters 29 Season and Canopy type 29 Physical Environment 31 Fish responses to variation in headwater drift subsidies 32 Bioenergetics models : 32 DISCUSSION 34 Terrestrial Inputs to Headwaters 34 Subsidy export from Headwaters 35 Fish responses to variation in headwater drift subsidies 36 Bioenergetics modeling 37 Estimating the contribution of headwater systems to downstream fish populations.... 39 Land Use 41 Conclusions 42 FIGURES AND TABLES FROM RESULTS 43 REFERENCES 57 CHAPTER 3: LOOKING FORWARD 66 Implications 66 Suggestions for further research 68 References 70 APPENDICES 71 Appendix A 71 Appendix B 72 Appendix C 73 LIST OF TABLES Table 1. Comparison of physical stream characteristics in the two stream basin study areas. Values reported are means with ranges in brackets. M K R F = Malcolm Knapp Research Forest 43 Table 2. General linear mixed model output for comparisons of drift and terrestrial abundance and biomass. Comparisons were made by season, canopy type and invertebrate origin in the drift (aquatic or terrestrial). A l l second-order interactions were not significant 44 Table 3. Back-transformed least-squares means estimates of abundance and biomass based on season, canopy and origin of invertebrates in the drift. A l l estimates are based on general linear mixed models with restricted estimate maximum likelihood (REML) estimation 45 Table 4. Different estimation methods of fish potentially supported by subsidies from an average 50 m reach of headwater stream. Estimates are based on canopy type and season. Numbers reported are averages 46 Table 5. Abundance and biomass of invertebrate terrestrial input and drift from headwater streams in the Pacific Northwest. Terrestrial input measured in 2 3 mg/m /day. Drift measurements in mg/m 47 Appendix A. Prospective and retrospective sensitivity analysis input parameters. Average parameter values were used to create a baseline model for comparison with changed parameter values. Prospective sensitivity analysis used parameter values increased by 10% of average value. Retrospective sensitivity analysis relied on the range of each parameter, increasing the value by 10% of the range. This type of analysis therefore allows for incorporation of the variability of each parameter in the model and provides a more conservative means for comparing variable change 71 Appendix B. Prospective and retrospective sensitivity analysis output (Fish Bioenergetics 3.0, Madison, WI). Prospective sensitivity analysis used parameter values increased by 10% of average value. Retrospective sensitivity analysis relied on the range of each parameter, increasing the value by 10% of the range. This type of analysis therefore allows for incorporation of the variability of each parameter in the model and provides a more conservative means for comparing variable change. Model output was generated by sequentially modifying each parameter and simulating the model for forty days. Values reported are estimated final mass of individual fish using parameter values reported in Appendix A 72 Appendix C. Prospective and retrospective sensitivity analysis of bioenergetics model parameters influencing predicted trout final masses (Fish Bioenergetics 3.0, v i Madison, WI). Prospective sensitivity analysis used parameter values increased by 10% of average value. Retrospective sensitivity analysis relied on the range of each parameter, increasing the value by 10% of the range. This type of analysis therefore allows for incorporation of the variability of each parameter in the model and provides a more conservative means for comparing variable change. Model output was generated by sequentially modifying each parameter and simulating the model for forty days. Values reported are percent change in growth due to a change in parameter value 73 vii LIST OF FIGURES Figure 1. Map of southwestern British Columbia with Malcolm Knapp Research Forest and the Chilliwack River Basin. Stars indicate the location of stream study areas.. 48 Figure 2. (a) Abundance and (b) biomass of terrestrial invertebrate inputs per week throughout summer 2006 in two headwater streams of broadleaf and coniferous canopy. Regression slopes are not significantly different from one another but are significantly different from zero. Week 1 = May 19-24, 2007, Week 15 = August 23 -September 1, 2007 49 Figure 3. (a) Abundance and (b) biomass of invertebrates in the drift by season, canopy and origin (least-square means ± 1SE). Bars with a '* ' are significantly different at the 0.05 level 50 Figure 4. (a) Abundance and (b) biomass of terrestrial invertebrates inputs by season and canopy (least-square means ± 1SE). Bars with a '* ' are significantly different at the 0.05 level : 51 Figure 5. Relative growth rates of cutthroat trout in Mayfly Creek channels based on experimental addition of terrestrial invertebrates. Regression line based on response per channel (n = 10) (see text for treatments included in the analysis). Individual fish (n = 30) plotted. Food addition at zero includes baseline drift in channels (0.038 g/channel/day on average). Fish in highest food addition (input rate = 0.96 g/channel/day) are receiving maximum daily ration 52 Figure 6. Average daily temperature per two-week sampling period in Mayfly Creek experimental channels. Average temperature from each sampling period was significantly different from all other periods. Lower and upper edges of box represent first and third quartiles. Middle line in box represents median. Lines extending outside of box represent smallest and largest non-outlier observations. Dots outside box represent outliers 53 Figure 7. Changes in relative growth rate across food treatments per time. Regression slopes are not significantly different from one another. Food addition at zero includes baseline drift in channels (0.038 g/channel/day on average). Fish in highest food addition (input rate = 0.96 g/channel/day) are receiving maximum daily ration. 54 Figure 8 (a) Prospective (conventional) and (b) retrospective (relative) sensitivity analysis output. At low prey densities, temperature and initial mass produce the largest changes in final mass. However, at high prey densities, changes in proportion of aquatic invertebrates produces the greatest change. Values plotted are percent change in final mass per food ration level over the 40 days of the simulation for each parameter. Estimates for prospective analysis are derived by calculating percent v i i i change in final mass from an initial model of average values to one where each parameter is increased individually by 10% of its average value. Estimates for retrospective analysis are derived by calculating percent change in final mass from an initial model of average values to one where each parameter is increased individually by 10% of its range. Terrestrial and aquatic invertebrate values were changed by increasing the proportion of each invertebrate category (while decreasing the proportion of the other invertebrate category) 55 gure 9 (a) Relative growth rates from Mayfly Creek experiment and Wisconsin Bioenergetics model. Values plotted are individual relative growth rates from each fish in the experiment. Food addition at zero includes baseline drift in channels (0.038 g/channel/day on average). Fish in highest food addition (input rate = 0.96 g/channel/day) are receiving maximum daily ration, (b) Relative growth rates from Mayfly Creek experiment and Wisconsin Bioenergetics Model. Values plotted are channel averages ± 1SE. Food addition at zero includes baseline drift in channels (0.038 g/channel/day on average). Fish in highest food addition treatment (input rate = 0.96 g/channel/day) are receiving maximum daily ration 56 ix ACKNOWLEDGMENTS First and foremost, thank you to my supervisor, John Richardson, for support and guidance throughout this entire process. I would also like to thank my committee members, Scott Hinch and Jon Shurin, for valuable advice and assistance with all stages of my thesis. Members of the StARR lab were incredible in support, advice and camaraderie over the past two years. Finally, thank you to family and friends for sticking with me throughout. Indispensable field and lab assistance came from Ashlee Albright, Isabelle Deguise, Laurie Marczak, Melissa Hogg, Nancy Hofer, Phoebe Ho, Pina Viola, Shannon Turvey, Sylvia Wood, Wil l Gibson and Xavier Pinto. Ionut Aron assisted with coordination of research at the Malcolm Knapp Research Forest. Jerry Maedel helped with creation of GIS maps of study sites. Thank you to Dan Moore for loan of specialized equipment for canopy cover measures. A special thank you to Val LeMay and Rebecca Best for statistics advice. Trent Hoover and Laurie Marczak provided incredible assistance with manuscript editing and revision. This material is based upon work supported under a National Science Foundation Graduate Research Fellowship and a US-Canada Fulbright Fellowship. Research costs were covered by a Forest Sciences Program (BC) grant. Any opinions, findings, conclusions or recommendations expressed in this publication are those of the author and do not necessarily reflect the views of the National Science Foundation. CO-AUTHORSHIP STATEMENT Chapter 2 is being prepared for submission in a scholarly journal under the same title as given here. The co-author of this chapter is Dr. John Richardson. A . E. Reiss conducted the research, performed all of the data analysis, and wrote the manuscript. The co-author helped design the study and improve the manuscript. CHAPTER 1: ECOSYSTEM SUBSIDIES AND SERVICES -VALUING RESOURCE FLOW Movement of materials between ecosystems links donor and recipient habitats while augmenting resources available to recipient communities (Bilby et al. 1996, Polis et al. 1997, Knight et al. 2005). The study of ecosystems as bounded entities has led to numerous advances in all fields of ecology. Recently, recognition of the magnitude and extent of cross-ecosystem resource flow, or subsidies, has broadened our perspective from the local to larger scales. In the process, the tangible lines defining ecosystems have become blurred. Subsidies are typically defined as any material or energy that moves across an ecosystem boundary and elicits a response in the recipient community (Polis et al. 1997). When we speak of subsidies therefore, we often use the terms donor and recipient ecosystem to distinguish contributing and receiving systems, respectively. In this introduction, I discuss three important components of subsidies: the ubiquity of subsidies, the consequences of subsidies and the value of services provided by subsidies. Subsidies are everywhere Cross-ecosystem flows of material or energy, or ecosystem subsidies, form a dynamic link between different habitats. The movement of material from terrestrial to aquatic systems is one of the most obvious examples of a subsidy (Cummins and Klug 1979). Invertebrates or leaves falling from the riparian canopy into a water body are an important flux of material in aquatic habitats (Caraco and Cole 2004, Nakano et al. 1999, Baxter et al. 2005). However, there are many examples in the literature of ecosystem subsidies between all types of systems. Marine-derived nutrients from spawning salmon 1 carcasses are deposited in freshwater systems (Bilby et al. 1996, Cederholm et al. 1999, Zhang et al. 2003) and then further transported into riparian forests by bears (Hilderbrand et al. 1999). This movement of material is an example of subsidy movement from nutrient-rich to nutrient-poor systems. In oceanic environments, the death and subsequent sinking of a whale can provide an enormous subsidy to depauperate ocean floors (Smith and Baco 2003). Similarly, in urban environments, bird populations have been found to be an important vector for the transport of nutrients into fragmented forest habitat through their feces (Fujita and Koike 2006). Sometimes, the presence of subsidies can drive movement patterns of consumers. In coastal California, coyote abundance and viability was linked to near-shore marine environments and, more specifically, to marine food resources (Rose and Polis 1998). Recently, flow of resources from Ashless headwaters into downstream fish-bearing reaches has also been considered a source of subsidy to downstream fish populations (Wipfli and Gregovich 2002). Fish living in downstream reaches feed upon drifting invertebrates present in the fish-bearing reach and presumably include those exported from Ashless headwaters. However, in this situation material is moving within a stream network, fish predators are limited to the downstream reaches. Most often, geographic limitation is either due to presence of fish barriers in the stream or simply the small size of headwaters resulting in insufficient volume of water to support fish. Given that streams in the Pacific Northwest are nutrient poor, any supplementation from external systems (i.e. upstream) stands to benefit recipient consumers. Due to the apparent ubiquitous nature of cross-ecosystem resource flows, the dynamics of this potential recipient 2 ecosystem must be explored in order to assess the relative contributions of these subsidies to recipient populations. While it may no longer be a surprise that ecosystems are connected, the extent and magnitude of these linkages likely varies between systems. Consequences of subsidies Although recognized and documented extensively across systems in recent years, the magnitude and consequences of cross-ecosystem flows remain poorly understood. Fundamentally, the type of subsidy is important, as, in order to be considered a subsidy, it must have a utility in the life history of direct or indirect consumers in the recipient community. It has also been posited that the magnitude of consumer response to subsidies should be a function of the net primary productivity (NPP) in the donor and recipient habitats (Persson et al. 1996, Polis et al. 1997, Marczak et al. 2007). Specifically, donor systems with high NPP should produce the largest recipient community responses when flowing into depauperate systems, while recipient systems that have low NPP, or are in a sense resource limited, should also produce the largest consumer response. In other words, a system with limited resources stands to benefit the most from supplementation. Most subsidy research therefore has focused on donor systems with high productivity and recipient systems with low productivity, such as movement from terrestrial to aquatic systems. Not surprisingly, we often see consumer responses to subsidies in these systems (Pace et al. 2004). While these trends in subsidy to production relationships have been observed in many studies, a recent meta-analysis on subsidy influence has demonstrated that they do not hold true overall in subsidy research (Marczak et al. 2007). The lack of accord of 3 subsidy-to-production-relationships with predicted theory might be due to the limited number of studies of subsidies moving from low NPP systems into high NPP systems. A recent study of subsidy influence of emerging adult aquatic insects (low NPP donor system) on riparian spider communities (high NPP recipient system) demonstrated just that (Marczak and Richardson 2007). Presumably, therefore, systems with high NPP stand to benefit from subsidies as well. An alternative explanation for the magnitude of a consumer response to a subsidy is that the ratio of subsidy to recipient, or ambient, resource within the recipient habitat is perhaps more important than the actual NPP within either system (Marczak et al. 2007). In particular, therefore, we should be able to predict the magnitude of response to a subsidy through examining the relationships between donor and recipient ecosystems. Validation of these subsidy-to-consumer-response relationships is most easily achieved by measuring consumer response directly. In the case of subsidies from Ashless, headwater streams to downstream, fish-bearing reaches, we explore a new aspect of subsidy to production relationships: movement of resources from one low productivity system into another. At the same time, given that headwater streams are much more numerous than fish-bearing reaches in an average watershed (Gomi et al. 2002), we would expect to see consumer responses commensurate with high subsidy to ambient resource ratios. Ecosystem services provided by subsidies The concept of ecosystem services has received much attention since its conception as a mechanism for conserving valuable resources (Daily 1997, Millennium Ecosystem 4 Assessment 2005). The premise behind the quantification of ecosystem services is to assign an economic value to non-resource-based products in the landscape (Kremen 2005, Kremen and Ostfeld 2005). Specifically, this approach values "support resources," or those components of an ecosystem that contribute to resources that are beneficial to human communities and populations. Classic examples of ecosystem services include pollination services offered to coffee plantations from neighboring forests (Ricketts et al. 2004) and water purification services provided by soil microbial communities (Graham and Smith 2004). Not only do ecosystem services provide a means by which to attribute value to non-resource-based ecosystem components, but they also generate novel ways to incorporate conservation planning into our management regimes (Chan et al. 2007). Applying the concept of ecosystem services to ecosystem subsidies extends the value of subsidies to both direct and indirect consumers. Given that fish stocks are considered a provisioning ecosystem service (Daily 1997), any subsidy that promotes or supports fish production should also be considered a part of that ecosystem service. While forested watersheds in and of themselves can provide ecosystem services, such as clean drinking water or hydroelectric power (Guo et al. 2000), invertebrate drift, which increases fish prey resources, can also be considered an ecosystem service. Ecosystem subsidies therefore have the potential to be valued as an ecosystem service to recipient ecosystems when the response occurs in human-valued resources. This current study provides a means of evaluating both subsidies and ecosystem services. Research Questions and Hypotheses Given our current knowledge of subsidies, I asked three general questions in this thesis: 5 1. How do seasonal and land-use induced changes influence the rate of subsidy movement from Ashless headwater streams to downstream fish-bearing reaches? 2. How do downstream consumer populations respond to variation in this subsidy and to what extent does this response fit expected model predictions? 3. What is the magnitude of the ecosystem service or value of subsidies from headwaters to both downstream consumers and recipient ecosystems as a whole? I investigated the relationship between seasonal variation in downstream invertebrate transport and upstream forest management. Specifically, by looking at early and late successional forest types, I contrasted the influences of different forest management practices on insect communities present in the water column (drifting invertebrates). In the Pacific Northwest, forest harvesting can often initiate a successional dynamic of red alder (Alnus rubra) communities, which transition into coniferous ones. I tested the hypothesis that early successional forests have higher rates of secondary production in selected aquatic and terrestrial macroinvertebrate species than late successional forests, which have transitioned into Western red-cedar (Thuja plicata), and Western hemlock (Tsuga heterophylla). These predicted differences could be due to increased light in streams of early successional stands as well as higher nutritional quality leaf litter contributions from the alder canopy (Cummins et al. 1989, Price et al. 2003, Banks 2005). Seasonally, I predicted that the highest export of headwater invertebrates would occur in the spring with diminishing resource flow throughout the summer and early fall. This prediction is based on diminishing water flows in headwater streams throughout the summer as well as observed declines in abundances of aquatic invertebrates, the most numerous component of drift samples, in streams throughout the world (Baxter et al. 2005). I propose that changes in insect communities with season or canopy type would be 6 transmitted through the watershed network, augmenting the rates of invertebrates transported in stream flow and thus available as a potential food subsidy to downstream fish populations. Further, I tested the hypothesis that headwater subsidies are consumed at a rate relative to their abundance by downstream cutthroat trout {Oncorhynchus clarki Richardson) populations, increasing trout growth rates and providing evidence for food limitation in downstream communities. In doing so, I address the question of what impact variability in these resources has on fish growth and predicted that increased subsidies would result in greater fish growth (Mason 1976). In testing this hypothesis, I determined the shape of their growth response curve and validated my results with a bioenergetics model. I hypothesized that differences in growth rate between experimental and model outcomes could be explained by alternative pressures on fish feeding and growth. Through knowledge of variation in headwater inputs as assessed in the part of my study focused on headwater exports, I determined the extent to which variation in the amounts and timing of headwater food resources influence cutthroat trout growth. 7 References Banks, J. L. 2005. Influences of clearcut logging on macroinvertebrates in perennial and intermittent headwaters of the central Oregon coast range. MSc. thesis. Dept. of Fisheries Science. Oregon State University, Corvallis, Oregon, USA. 188pp. Baxter, C. V. , K. D. Fausch, and W. C. Saunders. 2005. Tangled webs: Reciprocal flows of invertebrate prey link streams and riparian zones. Freshwater Biology 50:201-220. Bilby, R. E., B. R. Fransen, and P. A. Bisson. 1996. Incorporation of nitrogen and carbon from spawning coho salmon into the trophic system of small streams: Evidence from stable isotopes. Canadian Journal of Fisheries and Aquatic Sciences 53:164-173. Caraco, N . , and J. Cole. 2004. When terrestrial organic matter is sent down the river: The importance of allochthonous carbon inputs to the metabolism of lakes and rivers. Pages 301-316 in G. A . Polis, M . E. Power, and G. R. Huxel, editors. Food webs at the landscape level. University of Chicago Press, Chicago, Illinois, USA. Cederholm, C. J., M . D. Kunze, T. Murota, and A. Sibatani. 1999. Pacific salmon carcasses: Essential contributions of nutrients and energy for aquatic and terrestrial ecosystems. Fisheries 24:6-15. Chan, K. M . A. , R. M . Pringle, J. Ranganatran, C. L. Boggs, Y . L. Chan, P. R. Ehrlich, P. K. Haff, N . E. Heller, K. Al-Krafaji, and D. P. Macmynowski. 2007. Conservation planning for ecosystem services. Conservation Biology 21:59-68. Cummins, K. W., and M . J. Klug. 1979. Feeding ecology of stream invertebrates. Annual Review of Ecology and Systematics 10:147-172. Cummins, K. W., M . A. Wilzbach, D. M . Gates, J. B. Perry, and W. B. Taliferro. 1989. Shredder and riparian vegetation - Leaf litter that falls into streams influences communities of stream invertebrates. Bioscience 39:24-30. Daily, G.C., Ed. 1997. Nature's services: Societal dependence on natural ecosystems, Island Press, Washington, DC. Fujita, M . , and F. Koike. 2006. Birds transport nutrients to fragmented forests in an urban landscape. Ecological Applications 17:648-654. Graham, D. W., and V. H. Smith. 2004. Designed ecosystem services: Application of ecological principles in wastewater treatment engineering. Frontiers in Ecology and the Environments 2:199-206. Gomi, T., R. C. Sidle, and J. S. Richardson. 2002. Understanding processes and downstream linkages of headwater systems. Bioscience 52:905-916. 8 Guo, Z., X . Xiao, and D. L i . 2000. An assessment of ecosystem services: Water flow regulation and hydroelectric power production. Ecological Applications 10:925-936. Hilderbrand, G. V. , T. A . Hanley, C. T. Robbins, and C. C. Schwartz. 1999. Role of brown bears (Ursus arctos) in the flow of marine nitrogen into a terrestrial ecosystem. Oecologia 121:546-550. Knight, T. M . , M . W. McCoy, J. M . Chase, K. A. McCoy, and R. D. Holt. 2005. Trophic cascades across ecosystems. Nature 437:880-883. Kremen, C. 2005. Managing ecosystem services: What do we need to know about their ecology? Ecology Letters 8:468-479. Kremen, C , and R. S. Ostfeld. 2005. A call to ecologists: Measuring, analyzing, and managing ecosystem services. Frontiers in Ecology and the Environment 3:540-548. Marczak, L . B., and J. S. Richardson. 2007. Spider and subsidies: Results from the riparian zone of a coastal temperate rainforest. Journal of Animal Ecology 76:687-694. Marczak, L . B., R. M . Thompson, and J. S. Richardson. 2007. Meta-Analysis: Trophic level, habitat, and productivity shape the food web effects of resource subsidies. Ecology 88:140-148. Mason, J. C. 1976. Response of underyearling coho salmon to supplemental feeding in a natural stream. Journal of Wildlife Management 40:775-788. Millennium Ecosystem Assessment. 2005. Living beyond our means: Natural assets and human well-being. Island Press, Washington. Nakano, S., Y. Kawaguchi, Y . Taniguchi, H. Miyasaka, Y . Shibata, H. Urabe, and N . Kuhara. 1999. Selective foraging on terrestrial invertebrates by rainbow trout in a forested headwater stream in northern Japan. Ecological Research 14:351-360. Pace, M . L. , J. J. Cole, S. R. Carpenter, J. F. Kitchell, J. R. Hodgson, M . C. Van de Bogert, D. L. Bade, E. S. Kritzberg, and D. Bastviken. 2004. Whole-lake carbon-13 additions reveal terrestrial support of aquatic food webs. Nature 427:240-243. Persson, L. , J. Bengtsson, B. A. Menge, and M . E. Power. 1996. Productivity and consumer regulation - Concepts, patterns and mechanisms. Pages 396-434 in G. A. Polis and K. O. Winemiller, editors. Food webs: Integration of patterns and dynamics. Chapman & Hall, New York. Polis, G. A. , W. B. Anderson, and R. D. Holt. 1997. Toward an integration of landscape and food web ecology: The dynamics of spatially subsidized food webs. Annual Review of Ecology and Systematics 28:289-316. 9 Price, K., A . Suski, J. McGarvie, B. Beasley, and J. S. Richardson. 2003. Communities of aquatic insects of old-growth and clearcut coastal headwater streams of varying flow persistence. Canadian Journal of Forest Research 33:1416-1432. Ricketts, T. H., G. C. Daily, P. R. Ehrlich, and C. D. Michener. 2004. Economic value of tropical forest to coffee production. Proceedings of the National Academy of Sciences of the United States of America 101:12579-12582. Rose, M . D., and G. A . Polis. 1998. The distribution and abundance of coyotes: The effects of allochthonous food subsidies from the sea. Ecology 79:998-1007. Smith, C. R., and A . R. Baco. 2003. Ecology of whale falls at the deep-sea floor. Oceanography and Marine Biology 41:311-354. Wipfli, M . S., and D. P. Gregovich. 2002. Export of invertebrates and detritus from Ashless headwater streams in southeastern Alaska: Implications for downstream salmonid production. Freshwater Biology 47:957-969. Zhang, Y. X , and J. N . Negishi, J. S. Richardson, and R. Kolodziejczyk. 2003. Impacts of marine-derived nutrients on stream ecosystem functioning. Proceedings of the Royal Society of London Series B-Biological Sciences 270:2117-2123. 10 CHAPTER 2: LINKING HEADWATER SUBSIDIES TO DOWNSTREAM CONSUMERS1 INTRODUCTION Nutrients, detritus and prey move between ecosystems. These movements of resources, or subsidies, from contributing systems can increase receiving, or recipient, habitat productivity and alter consumer-resource dynamics (Bilby et al. 1996, Polis et al. 1997, Knight et al. 2005). Examples of these transfers can be found in most ecosystem types including windborne subsidies of insects and detritus on land (Polis et al. 1997), organic matter transport from upstream to downstream stream reaches (Vannote et al. 1980), and movement of leaf litter and terrestrial invertebrates into aquatic systems (Caraco and Cole 2004, Nakano et al. 1999, Baxter et al. 2005). Recently, the movement of invertebrates and detritus from Ashless headwater streams has been considered as a subsidy for downstream fish populations (Wipfli 2005). The apparently ubiquitous nature of between-ecosystem subsidies has reinforced the notion that ecosystems could be more open than previously conceived (Polis et al. 1997, Palumbi 2003). Most researchers have assumed that the magnitude of consumer responses to subsidies was likely dependent on a number of factors, including net primary productivity in both donor and recipient habitats (Polis and Hurd 1996). The extent of these linkages, however, has been shown in a recent meta-analysis to be driven instead by the ratio of subsidy to recipient system, or ambient, resources (Marczak et al. 2007). Knowledge of how various subsidies differ in their value to organisms in recipient systems is crucial to understanding how adjacent 1 A version of this chapter will be submitted for publication. A. E. Reiss and J. S. Richardson. The Invertebrate Connection: tracing the value of food subsidies from Ashless headwaters on downstream Ash-bearing reaches. Canadian Journal of Fisheries and Aquatic Sciences. 11 ecosystems are connected. By predicting how cross-ecosystem flows of resources alter the production of a human-valued resource (i.e., fish), we also determine the ecosystem service or value of subsidies to direct and indirect consumers (Daily 1997). Stream ecologists have long recognized the important role of terrestrial inputs for understanding stream functioning (Egglishaw 1964, Mason and MacDonald 1982). This is particularly true in the case of small headwater streams. Secondary production in these streams is often driven by inputs of terrestrial leaf litter, while in situ production is frequently an unimportant fraction of total system productivity (Angelier 2003). Connectivity between terrestrial and riparian habitats results in large flows of materials into aquatic communities, subsequently supporting higher levels of secondary productivity than would occur from primary production alone (Wallace et al. 1997, Nakano and Murakami 2001, Gomi et al. 2002, Wipfli et al. 2007). Terrestrial inputs also take the form of accidental inputs of invertebrates falling from the riparian canopy onto the surface of streams. These terrestrial invertebrates combine with drifting benthic invertebrates as potential food for vertebrate predators such as fish. Headwater systems that lack fish nevertheless have the potential to support downstream fish populations through the provision of drifting aquatic and terrestrial invertebrates. The subsidy of terrestrial invertebrate prey from headwater systems can be substantial (Wallace et al. 1997, Nakano et al. 1999), although the movement rate of these invertebrate subsidies into downstream systems has yet to be investigated on a large scale (Svendsen et al. 2004). Headwater systems are important conduits for the transport of 12 water, nutrients and material into downstream watershed networks (Vannote et al. 1980). For downstream salmonid populations, exports of invertebrates from these headwaters in Alaska are a significant potential food source (Wipfli and Gregovich 2002). In addition to invertebrate exports, organic matter is also exported from headwater streams to downstream reaches. Organic matter, while not directly consumed by fish, can increase the production of benthic invertebrates downstream, thus indirectly supporting fish populations (Bilby and Bisson 1992). Given that headwater streams can make up 80% of stream length in an average watershed (Gomi et al. 2002) and that invertebrates have been reported to drift on average up to 50-100 m (Gerking 1994), the ratio of headwater subsidy to downstream ambient resource has the potential to be quite high per unit area of downstream reach. Cumulatively we would expect this subsidy to elicit a broad response in consumers downstream (MacDonald 2000). However, the degree to which downstream consumer populations utilize upstream inputs has not yet been demonstrated experimentally. Although a large number of studies have detailed the movement of terrestrial subsidies to headwaters streams, relatively few have documented the contributions of Ashless headwater streams to downstream fish-bearing reaches. Still fewer have focused on the potential dependence on these resources by downstream fish populations, while none has measured actual use. Salmonid fish diets during freshwater life-stages consist mostly of terrestrial and aquatic invertebrates drifting in the water column (Allan et al. 2003). Fish biomass and abundance have been strongly linked to drift densities in a number of studies (Allan 1981, Dedual and Collier 1995, Johansen et al. 2005). However, prey availability 13 is often a limiting factor in stream fish populations. For example, cutthroat trout are often food limited, especially during the late summer months when increased temperatures create additional stresses (Mason 1976, Boss and Richardson 2002, Bacon et al. 2005). While drift occurs throughout the day, it peaks at night (Gries et al. 1997), when feeding is typically limited due to low visibility (Allan 1981, Angradi and Griffiths 1989). Further, the seasonal relationship between food availability and growth is not always synchronized. Specifically, in the Pacific Northwest, winter is a time of high flow and mass movement of materials (Karlsson et al. 2005), while the majority of growth in freshwater fish occurs between spring and fall (Hunt 1969, Boss 1999). Additional prey resources provided from headwater flow during the primary growth period might therefore play a critical role in trout growth. Although it remains to be seen definitively if size affects overwinter survivorship in coastal streams in British Columbia (S. Boss, University of British Columbia - unpublished data), this relationship has been well established in many other parts of the world (Hunt 1969, Smith and Griffith 1994). Further, we know that size does relate directly to egg production and reproductive rates in subsequent seasons (Moyle and Cech 1988). Therefore, the benefits of greater access to food during important periods of growth may be important for future fitness (Boss and Richardson 2002). A number of biotic and abiotic constraints have been found to limit fish growth, even during critical growth periods. For example, types of food resources for fish vary throughout the spring to fall. While spring is a time of high aquatic invertebrate consumption, in summer, fish rely more upon terrestrial invertebrates (Nakano and 14 Murakami 2001, Wipfli 1997). This variation in diet is largely due to the size, availability and visibility of different types of invertebrates (Ware 1972, Bisson 1978, Wilzbach et al. 1986, Angradi and Griffith 1989). Consumption of different proportions of terrestrial and aquatic prey by fish also occurs between streams with different riparian canopy types (Romero et al. 2005) or different light levels (Wilzbach et al. 2005). Stream temperature, body fat reserves and season play a role in determining fish growth (Bacon et al. 2005), as does channel depth (Harvey et al. 2005). It is difficult to simultaneously assess with controlled experiments how these factors can affect feeding behaviour and growth. I will use a bioenergetics modeling approach to help evaluate this issue and provide insights into how resource utilization constrains the growth of fish in naturally-occurring fish populations. In forested areas of the Pacific Northwest, small headwater streams receive little or no protection under most legislated guidelines, particularly when fish are absent (Moore and Richardson 2003). The result is that forest harvesting can directly alter the rates and quantities of terrestrial subsidies to these small streams. Further, harvesting can result in increases in light penetration, water temperature and sediment load to streams (Hassan et al. 2005, Moore et al. 2005), while physical features of the streams such as channel width (Sweeney et al. 2004), amount of large woody debris (LWD) (Murphy and Koski 1989) and bed roughness can also be severely modified. These changes might directly influence fish populations by manipulating habitat and might also alter water turbulence and flow patterns, which can impact drifting invertebrate communities (Gayraud et al. 2002). In regeneration after disturbance, streamside forest canopy composition can be altered, 15 modifying leaf litter and other organic inputs to streams (Price et al. 2003). A l l of these habitat changes alter the abundance and diversity of macroinvertebrate assemblages in streams (Cummins and Klug 1979, Hawkins et al. 1982, Ward and Stanford 1982, Cummins et al. 1989, Sweeney 1993). Further, differing land use practices have been shown to influence the abundance and biomass of terrestrial invertebrates falling into streams in New Zealand (Edwards and Huryn 1996) and Oregon (Meehan 1996, Romero et al. 2005). Therefore, logging adjacent to headwater streams is expected to have subsequent indirect impacts on recipient fish populations and riparian food webs as a whole. Forest harvesting converts older coniferous stands to areas frequently dominated by deciduous trees such as red alder (Alnus rubra), particularly along riparian corridors. For small streams that receive little or no mandated reserve strips, this canopy conversion can be dramatic. For example, the overall annual contribution of invertebrates from mature, coniferous forested headwaters in Alaska has been found to provide adequate biomass to potentially support 100-2000 young of the year salmonids per kilometer of fish-bearing stream (Wipfli 2005). Conversely, headwaters with young alder canopy in Alaska have been found to export roughly three times more biomass than young coniferous ones and 1.3 times more biomass than old growth coniferous ones (Piccolo and Wipfli 2002, Wipfli and Musslewhite 2004). Given that changes in riparian canopies alter terrestrial invertebrate inputs, harvesting may alter the resources available to fish populations over longer time scales. Increasing our knowledge of how different forest successional stages 16 contribute to stream ecosystems will help us to understand the ecological dynamics of long-term shifts in resource availability. I had four objectives in this chapter. First, I characterized the temporal variability of terrestrial inputs to streams with differing canopy types throughout the summer growth period of fish (roughly May to September). Secondly, I determined how invertebrates and subsidies of organic matter from Ashless headwaters varied between spring and summer and between coniferous and deciduous forest canopies. Quantification of the materials exported by the streams in this study permits a geographically-explicit comparison with the amounts of similar resources exported by headwaters in other parts of the world. Third, I experimentally assessed the growth of cutthroat trout (Oncorhynchus clarki) in response to a range of potential headwater-supplied invertebrate subsidies and compared these outcomes to the growth predictions of a commonly-used bioenergetics model. Finally, I utilized the information from these results to calculate the value of ecosystem service (i.e. food resource) provided by Ashless headwaters to fish production. By determining the extent to which a model salmonid, cutthroat trout, accrues fitness benefits from different drift densities of terrestrial invertebrate inputs, I examined overall salmonid response to variation in drift. Establishing a clear connection between salmonid fish fitness and headwater subsidies will contribute to a foundation for management of these currently underprotected areas. 17 METHODS Study Site This study was conducted on streams in the Malcolm Knapp Research Forest (MKRF) (49° 16' N , 122° 34'W) and the nearby Chilliwack River Basin (49°10.30' N , 121°56.30' W) (Figure 1), both located in the Coastal Western Hemlock biogeoclimatic zone of British Columbia, Canada. A total of 34 small headwater streams (of either broadleaf or coniferous canopy) were selected for maximal similarity in their physical characteristics (Table 1). Two of these streams were sampled weekly to assess long-term terrestrial input to streams. A l l 34 streams were used in a large-scale quantitative survey of invertebrate input and export into and from headwaters. Both study areas consist of forests of multiple age classes, from 1-year-old second growth to patches of 400+ year forest (Rood and Hamilton 1995). Terrestrial Inputs to Headwaters Season and Canopy Data Terrestrial invertebrate inputs to streams were quantified weekly throughout the summer. Pan trap samples (1 week) were collected at two streams (one of broadleaf and one of coniferous canopy) for fifteen weeks (May 19 th to September 1st). For pan trap samples, a clear plastic pan (20 x 28 cm) containing a small amount of water and a few drops of soap to act as a surfactant was placed at the edge of the active channel to collect invertebrates falling into the stream (see Allan et al. 2003). Samples were preserved in 7-18 8% formalin and sorted into two size classes for identification (250 fxm to 1 mm and >1 mm) based on feeding size preferences of trout (Bisson 1978). Individuals were identified to order and ashed to obtain biomass estimates. Samples were dried (at 60° for 24 h), weighed, burned to ash (at 550° for 2 h) and reweighed to determine ash-free dry mass (AFDM). Weekly pan trap samples were analyzed using an A N C O V A design with time as the covariate (PROC G L M , SAS 9.1, Cary, NC) to determine the influence of canopy type on terrestrial invertebrate inputs. Evaluation of the interaction between canopy and time allowed differences in slopes of treatments to be assessed. A type I error of 0.05 was used to assess significance in all models. Subsidy export from headwaters Season and Canopy Data To determine how subsidies of invertebrates and organic matter from headwaters varied with stream canopy type (broadleaf or coniferous), I measured invertebrate drift, terrestrial invertebrate inputs and several physical habitat parameters in 34 Ashless headwater streams. An equal number of streams with each canopy type (n = 17 of each type) were selected and these streams were distributed between the two study areas. A l l sampling was conducted in both spring (May) and summer (July). Streams were selected for consistency in size (1 to 3 m wetted width during initial survey in May), the absence of fish (determined by minnow trapping over 24 h and presence of fish barriers (Forest Practices Code of BC, 1998)) and canopy type (vegetation type had to be consistent for 19 100 m upstream of sampling site and comprise a minimum of 50% of the canopy of interest). Broadleaf riparian canopies selected were a minimum of 6 years old to ensure initial impacts of logging had subsided (Grant and Wolff 1991, Hartman et al. 1996, Marden et al. 2006). Coniferous forests were all second growth between the ages of 25-80 years. At each site, 24 h drift and pan trap samples were collected once during each of the two seasons. Sampling took place over ten days during each collection period with approximately nine streams sampled during each 24 h period. Day-to-day variation in samples was not expected to be greater than stream-to-stream variation (S. Nicol, University of British Columbia - unpublished data). To estimate sampling-day-variation, two specific streams were sampled during each 24 h sampling period. Invertebrates were captured using 250 um drift nets (15 x 25 cm) (Whitehouse Plankton Nets, Kamloops, BC). Stream velocity measurements were made using a Swoffer 2100 wand (Swoffer Instruments, Inc. Seattle, WA) at the mouth of the drift net at the beginning and end of each sample period in order to estimate the amount of water passing through each net. Pan trap samples were collected as described above in order to obtain estimates of abundance. Drift and pan samples were preserved and sorted as described above. Individuals in drift and pan samples were identified to order and then invertebrates in drift samples were characterized as either of terrestrial or aquatic origin based on published larval life history details (Bland et al. 1978, Merritt and Cummins 1996). After identification, 20 invertebrate biomass for each sample was estimated as described above. Organic matter collected in the 24 h drift net was sorted into two size classes. I designate coarse particulate organic matter (CPOM) as >500 \xm and fine particulate organic matter (FPOM) as >250 urn but less than <500 urn. Organic matter biomass was obtained through A F D M estimates as described above. Samples were converted to estimate/m2/day for pan traps through pan trap area, and estimate/m3 for drift samples using trap size and stream velocity measurements. The abundance and biomass of drifting invertebrates were each analyzed using a general linear mixed model (PROC MIXED, SAS 9.1, Cary, NC) with canopy cover (broadleaf vs. coniferous), season (spring vs. summer) and origin (terrestrial vs. aquatic invertebrates) as fixed factors. Abundance and biomass of terrestrial input was similarly analyzed using only canopy cover and season. Study area (MKRF vs. Chilliwack River Basin) was a random factor in the analysis. I used an identical approach to analyze the effects of canopy cover, season, and study area on the amount of C P O M and FPOM in drift samples. Abundance and biomass data were logio+1 transformed to successfully normalize variance. Tukey-Kramer HSD post-hoc tests were used to test for significant differences amongst treatments (Kutner et al. 2005). Specification of a compound symmetry covariance structure accounted for the non-independence of seasonal data. A Kenward-Rogers approximation correctly estimated the denominator degrees of freedom (Littell 2002). The influence of study area (i.e. block) was tested in mixed models by comparing the difference in -2 residual log likelihood of a full and reduced model (one where block as a random factor was dropped) to a X2 distribution. Best-fit models were selected based upon lowest AIC values. 21 Physical Environment Habitat parameters were estimated for each stream during each round of invertebrate sampling (May and July). These parameters included: gradient (Suunto clinometer Espoo, Finland), wetted width, largest particle, channel depth, % large woody debris (see Hogan 1996 for particulars on these four measurements), % open canopy (Nikon Coolpix 4500 Camera with Fisheye converter and Gap Light Analyzer 2.0, Burnaby, BC [Frazer 1999]), and stream temperature (HOBO® HO 1-001-01 water temperature data loggers recorded every 15 minutes, Pocasset, Massachusetts). Parameters were collected as potential covariates in the analysis (Griffith et al. 2001, Frady 2005). Principal components analysis using a correlation matrix was performed on environmental parameters to obtain site scores for bed roughness (PROC PRINCOMP, SAS 9.1, Cary NC). The site scores from the first principal component extracted were used as covariates in the analysis in an effort to account for differences in stream abiotic characteristics in all mixed models. Day to day variation in abundance and biomass of invertebrate samples was analyzed using a general linear mixed model as above (PROC MIXED) with canopy cover (broadleaf vs. coniferous), season (spring vs. summer), and sampling day (1-5) as fixed factors. The amount of large woody debris (LWD), percent open canopy cover, and stream temperatures (daily average, minimum and maximum values) were also each analyzed using an identical approach. Second order interactions were included in the mixed model where canopy type and season were fixed factors, while study area was 22 incorporated into the model as a random factor. L W D data were In transformed to successfully meet model assumptions. Fish responses to variation in headwater drift subsidies I examined cutthroat trout growth responses to variability in the abundance of drift subsidies using a series of fourteen experimental stream channels at Mayfly Creek in the M K R F (details on channels provided in Kiffney and Richardson 2001). Given exposure of fish to natural water, light and temperature regimes, and habitat features, these experimental channels can be considered a good representation of a natural system (Zimmerman and Vondracek 2006). Each channel was 12 m x 0.25 m and approximately 0.1 m deep. The channels were located on a floodplain next to Mayfly Creek, from where stream water was supplied to the channels by rerouting inflow through pipes from an upstream section. Water quality was consistent among channels (pH 7.4, dissolved oxygen 9.93 mg/L or 100%, turbidity 0.007 NTU). Channels were lined with black plastic covered by natural substrates (sand, gravel, pebble and cobble) and arranged in a series of riffles and pools. Bricks and cover boards were placed in each of the channels in order to provide consistent physical refuge for fish. Channel ends were secured with 0.4 cm 2 mesh and cleaned daily to remove buildup of detritus and excess food. Water temperature in each channel was recorded every 15 minutes with a water temperature data logger. Forty-two cutthroat trout (69-90 mm fork length) were collected by minnow trapping in nearby fish-bearing creeks. Trout were expected to be one-year-old using local size-age class reports (Boss 1999). Each fish was individually marked with a unique combination 23 of colored latex dye (Northwest Marine Technology, Shaw Island, WA) injected into the anal, caudal or pectoral fin. Initial biomass and fork length of each fish was measured. Three randomly selected fish were then introduced into each channel. The number of fish per channel was determined using reported natural fish densities (0.49 fish/m2) for headwater streams in the region (Young et al. 1999). The experiment was conducted over six weeks from June through mid-July. Fish mass (mg) and fork length (mm) were measured in the field for all individuals every two weeks using an Ohaus Scott //electronic field scale (Ohaus®, Pine Brook, New Jersey, USA). Seven food treatment levels were applied. Food was dispensed over the course of 12 hours using automatic, conveyor-belt style feeders (Boss and Richardson 2002). Freeze-dried shrimp (Hemimysis anomala) (Hagen®, Nutrafin basix mysis shrimp, Montreal, Quebec, Canada) were used as the food supplement as they provide a standardized food source that mimics floating terrestrial drift and have an energetic value that is comparable to natural invertebrate food (Cummins and Wuycheck 1971). Food treatments were randomized among channels and ranged from zero food addition to 9% per day of the initial average biomass (which was 5.05 g) of all fish in a channel. Food treatments were calculated so that the highest rates of food addition would exceed the maximum daily ration of 7% of a 5 g fish (Jobling, 1994), given average water temperatures of 12.3° in the channels throughout the experiment. Ambient drift in each channel was collected for 24 hours at the beginning and end of the six-week period as a potential covariate in the analysis. Relative fork length and relative growth rate were calculated as: 24 y = [(Final - Initial) / Initial] x 100 Regression analysis (PROC REG, SAS 9.1, Cary NC) was used to determine the effect of relative fork length on relative growth rate. A general linear mixed model (PROC MIXED) was used to determine the effects of food treatment on fish growth. Differences in temperature over time and the influence of temperature on relative growth rate were also tested (PROC G L M ) . Changes in relative growth rate over time in response to variable food addition were analyzed using an A N C O V A design with time as the covariate (PROC GLM) to determine differences in regression slopes. Bioenergetics Modeling The Wisconsin Bioenergetics Model (Fish Bioenergetics 3.0, Madison, WI) provides estimated growth outputs for cohorts of fish given known input parameters. I used this model to test the sensitivity of fish growth to variation in abiotic and biotic parameters (temperature, initial mass and changes in aquatic and terrestrial invertebrate drifting prey proportions). Variation in prey proportions adjusted the number of individual invertebrates of differing caloric content available to fish consumers. Second, I assessed the effectiveness of the model in predicting fish growth by comparing experimental and model outcomes. Differences between experimental and model data provide insight into the contribution of other factors aside from food availability on fish growth. Further, as model outcomes are based primarily on laboratory studies, comparisons with experimental results generated in natural settings highlight limitations of model predictions. 25 Bioenergetics modeling is based on the energy balance equation (Winberg 1956, Wootton 1990): C = P + R + F + TJ where C = food consumption, P = growth and reproduction (production), R = metabolism, F = fecal wastes, and U = excretory wastes (Gerking 1994). The Wisconsin Bioenergetics Model estimates growth (final mass) of a cohort based on a number of input parameters that help to model metabolism and consumption. I used temperature (measured in Mayfly Creek channels), proportions of different kinds of prey, i.e., terrestrial versus aquatic (from fish-bearing streams for sensitivity analysis and from Mayfly Creek fish growth experiment for model corroboration) as well as prey energy density (Cummins and Wuycheck 1971) as inputs to the model. Estimates of consumption were based on amount of prey present in the drift. Therefore, the model will overestimate actual fish growth given the fact that fish are unable to capture all invertebrates in the drift. However, growth response curves developed from experimental fish growth data were also based on food availability instead of food consumption. Therefore these two outputs should be readily comparable. To run the model, I simulated growth of individual fish as a cohort. Given similarities between growth of steelhead (Oncorhynchus mykiss) and cutthroat trout (Cartwright et al. 1998), I used physiological parameters for steelhead since they are not readily available for cutthroat trout. A l l model scenarios were "fit to consumption" without including spawning or mortality. Models were run for the same time and duration as the Mayfly 26 Creek fish growth experiment (June to mid-July). A l l other input parameters were kept at default settings (see Fish Bioenergetics 3.0, Madison, WI for complete list). Sensitivity of model predictions to the parameters was tested for important variables contributing to the growth of individual fish by both retrospective and prospective sensitivity analyses (Caswell 2000). Prospective (conventional) sensitivity analysis was performed by modifying each input parameter individually by 10% of its average value and comparing changes in predicted final mass of each individual as per Cross and Beissinger (2001). Retrospective (relative) sensitivity analysis was conductedby calculating the range of each variable for the experiment and then changing the value of each input parameter individually by 10% of its reported range. Retrospective analysis therefore allowed me to incorporate each parameter's variability (as a function of its range) in my analysis of parameter sensitivity. Average values and parameter ranges for stream temperature (from Mayfly Creek channels), proportion of both terrestrial versus aquatic invertebrates (based on terrestrial to aquatic invertebrate proportions in fish-bearing streams (Romero et al. 2005)), and initial mass (using fish growth data from Mayfly Creek experiments) were calculated and entered in the model and can be found in Appendix A . Prey energy densities for both the sensitivity analysis and modeled Mayfly Creek growth rates can also be found in Appendix A. For both types of analysis, I sequentially modified each parameter within a food ration level and allowed the model to estimate 40 days of fish growth (as in the Mayfly Creek experiments). Percent change in final mass was calculated as: y = [(Increased parameter value - Initial model) / Initial model] x 100 27 Predicted final masses of individual fish from Mayfly Creek were also generated with the Wisconsin Bioenergetics Model. Input parameters included average daily temperature, prey proportions, prey energy density (Cummins and Wuycheck 1971) as well as initial mass and ration of each individual modeled. A l l input values were those measured during the Mayfly Creek channel experiments. Predicted relative growth rates were then calculated as above (see Methods: Fish responses to variation in headwater subsidies) for comparison with actual measured relative growth rates. Differences between actual and predicted relative growth rates were calculated as in Liao et al. (2005). 28 RESULTS Terrestrial Inputs to Headwaters Abundance and biomass of terrestrial invertebrate input, as measured by weekly pan trap samples, were not significantly different between canopy types over time (Fi ;26 = 0.02, p = 0.89 and Fi ;23 = 0.03, p = 0.87, respectively). However, there was a significant increase in both abundance and biomass throughout the course of the summer in both streams (Fi, 26 = 4.84, p = 0.04 and F U 2 3 = 6.05, p = 0.02) (Figure 2a and 2b). Subsidy Export from Headwaters Season and Canopy type Overall, invertebrate abundance in the drift was 16% greater in broadleaf canopies than in coniferous canopies, although the difference was not significant, and was significantly higher (56%) in summer than in spring. Drift was composed of significantly more invertebrates of aquatic origin than terrestrial origin, with roughly 75% of all individuals sampled being of aquatic origin (Table 2, Figure 3a). Input rates of terrestrial invertebrates were 18% higher (although not significant) in broadleaf canopies. Inputs rates in both canopy types were significantly greater (1.1-times higher) in the summer (Table 2, Figure 4a). The interaction between canopy and season was not significant for abundance of drifting or terrestrial invertebrates (Fi;29.4 = 0.44, p = 0.51 and F i s 32.1 = 0.02, p = 0.88, respectively). Within the drift, the average A F D M of an aquatic invertebrate was 0.29 ± 0.14 mg, while that of a terrestrial invertebrate was 1.22 ± 0.47 29 mg. On average, drift samples consisted of 83% aquatic invertebrates and 17% terrestrial invertebrates. However, while there were more individual aquatic invertebrates than terrestrial in the drift, the mass proportion of terrestrial invertebrates (49%) in the drift was almost equal to that of aquatic invertebrates (51%). Significantly more invertebrate biomass (25%) was found in the drift in July than in May (Table 2, Figure 3b). Although biomass was 5% higher in the drift in broadleaf canopies relative to conifer forests, the difference was not significant. Biomass of terrestrial inputs was not significantly different based on treatment type, although inputs were 7% higher in summer and 9% higher in broadleaf canopies (Table 2, Figure 4b). The interaction between canopy and season was not significant for biomass of drifting invertebrates or for terrestrial invertebrate inputs (F ) ; 28.8 - 0.17, p = 0.68 and Fi^o = 0.0, p = 0.99, respectively). For the purposes of linking these data to the fish growth data, I also back-transformed all abundance and biomass results (Table 3). Inputs of C P O M were significantly greater (1.8 times higher) in summer than in spring (Fi,25.7 = 10.81, p = 0.003), but did not differ significantly between canopy types in this study (Fi(26.7 = 0.56, p = 0.46). C P O M output was estimated to be approximately 7 mg/m3 of stream water. F P O M did not vary significantly by season or canopy type (Fi_ 30.7 = 3.05, p = 0.09; Fi;29.4 = 0.36, p = 0.55). FPOM outputs were estimated at approximately 3.3 mg/m3 of stream water. The interaction of canopy and season was not significant for either C P O M or F P O M (F,,28.2 = 0.57, p = 0.46 and Fi , 3 0 .7 = 0.59, p - 0.44, respectively). 30 Therefore, as mg/m of stream water, there was 4.3-times more CPOM, and 2.0-times more FPOM moving downstream than drifting invertebrates. The first principal component of habitat parameters explained 32% of the variation in the data and was most explained by gradient. When measures of stream roughness (as assessed by PC eigenvalue scores) were included in the model, they were not significant for abundance or biomass of drifting or terrestrial invertebrates. Therefore, the covariate was dropped from the model (Engqvist 2005). Physical Environment Sampling day did not significantly influence the abundance of invertebrates in drift samples ( F 4 , n.i = 2.52, p = 0.10). Day-to-day variation in sampling was lower than that due to canopy type. Large woody debris (LWD) did not vary significantly by season (Fi, 29 = 1.80, p = 0.19) or canopy type (Fi, i = 0.06, p = 0.85). There was no significant interaction with L W D between canopy and season. Canopy was more open in all streams in spring (F 1,30.4 = 13.57, p = 0.001) and in streams with broadleaf canopies (F 1,29.7 = 6.40, p = 0.02). There was no significant interaction between canopy and season in terms of % canopy cover (Fi,3o.4 = 0.27, p = 0.61). Average, maximum, and minimum stream temperatures were higher in summer (F 1,28.9= 173.48, p < 0.0001; Fi , 3 0 . 8 = 86.54, p < 0.0001; and F i , 2 9 = 170.38, p < 0.0001, respectively), but not influenced by canopy type (Fi,29.7 = 0.02, p = 0.89; F i , 3 2 = 0.0, p = 0.98; and F,,29.1 = 0.0, p = 0.99, respectively). There was no significant interaction between canopy and season for temperature. Differences between the two river basins (MKRF and Chilliwack River Basin) did not 31 significantly influence the biomass of drifting or terrestrial invertebrates or the abundance of terrestrial input (all p > 0.66), but did have an effect on abundance of drifting invertebrates (p < 0.0001). Fish responses to variation in headwater drift subsidies Ambient drift in.the experimental channels averaged 0.49 ±0.18 mg/m A F D M or 0.038 g/channel/d. Relative growth rate increased significantly with relative fork length (r2 = 0.754, p < 0.0001). Furthermore, relative growth rate of fish increased as drifting prey inputs increased (Fi, 10.5= 5.32, p = 0.04). The two highest food treatments were removed from the regression due to high escape rates of fish from these channels. When these channels were removed from the analysis, the positive relation between food inputs and growth rates remained significant (F 1 ; g.38 = 12.06, p = 0.01) (Figure 5). Temperature increased significantly over the course of each two-week sampling period (F2,957= 247.59, p < 0.0001) (Figure 6). Overall, average daily temperatures during the six-week experiment averaged 12.3 °C and ranged from 8.6°Cto 17.1 °C. However, temperature increase from week one to week six did not significantly influence the relative growth rate at each sampling period (F2,36 = 2.60, p = 0.09). Relative growth rates did not vary significantly between each two-week sampling period ^2,33 = 1.07, p = 0.35) (Figure 7). Bioenergetics models Both conventional and retrospective sensitivity analysis determined that the proportion of terrestrial to aquatic prey was the most important contributor to fish growth in the model at high prey densities (Figure 8a and 8b). At lower prey densities, which are closer to 32 naturally observed levels in fish-bearing streams, temperature dominates growth. Predicted growth with increase in parameter value (either 10% of average parameter value or range of parameter value) and percent change in growth based on model output estimates for final mass are in Appendices B and C, respectively. Differences in predicted and actual relative growth rates were larger at low prey densities (Figure 9a). Overall, actual relative growth rates were higher than predicted ones. Observed and predicted relative growth rates were calculated to differ by 123% (Figure 9b). However, given large confidence intervals on both curves, the potential degree of overlap between the two lines is substantial. In particular, it is only at the lowest food rations that the curves diverge. 33 DISCUSSION We found that Ashless headwaters contribute substantial quantities of invertebrates and detritus to downstream fish-bearing reaches. Further, there were significant seasonal and suggestive canopy-related differences in these material exports. These results are consistent, both in terms of drift density and relative prey proportions, with findings related to headwater export in similar regions, most particularly in Alaska (Wipfli and Gregovich 2002, Wipfli and Musslewhite 2004). The growth of cutthroat trout in experimental channels increased with increased drift densities, suggesting a strong linkage between export rates and consumer response. Contrary to expectation, actual relative growth rates were higher than those predicted by the bioenergetics modeling, although only at low food densities. The difference might be due to the absence in the experimental channels of other energetic costs that are implicitly but not explicitly incorporated in the model, such as swimming behavior, that limit fish growth in the wild (Christiansen and Jobling 1990). Given these key findings, this study suggests that headwater subsidies can be altered by upstream land management practices and that availability of subsidy resources at critical growth periods for freshwater fish (spring-fall) has the potential to elicit a response in downstream consumers. Terrestrial Inputs to Headwaters Weekly measurements of terrestrial invertebrate inputs into the headwaters streams in this study demonstrated a significant increase in the number and biomass of invertebrates falling into streams from spring to fall. Many stream systems around the world, such as 34 those in Japan (Nakano and Murakami 2001) and the eastern United States (Romaniszyn et al. 2006) exhibit large fluxes in the availability of terrestrial and aquatic invertebrates on a seasonal basis. Although I did not collect measurements of fall and winter subsidies, fish size prior to winter has been found to be important in overwinter survival (Hunt 1969, Boss 1999). This suggests that the time period for our study might be the most critical in terms of fish growth. Coastal, fish-bearing streams receive not only pulsed subsidies such as those from spawning salmon (marine-derived nutrients) (Kline et al. 1997, Zhang et al. 2003), but also subsidies that fluctuate through time such as those from headwater catchments. Subsidy export from Headwaters While fish-bearing streams are generally moderately productive (Cushing and Allan 2001), much of the within-system productivity is based on resources derived from the adjacent terrestrial system. In coastal temperate rainforest in British Columbia, average invertebrate drift density in fish-bearing streams has been reported in the range of 0.23 mg/m3 (Rosenfeld and Boss 2001). Of in-stream prey biomass resources, 21% are of terrestrial origin (N. Romero, Oregon State University - pers. comm.). In comparison, I determined the invertebrate drift biomass in headwater streams to be 1.37 mg/m in streams adjacent to broadleaf riparian canopies and 1.30 mg/m in streams adjacent to coniferous riparian canopies at my study sites. Furthermore, terrestrial invertebrates comprised 49% of drift biomass resources in the headwater streams in this study. The higher proportion of terrestrial invertebrate biomass in Ashless headwater streams compared to values from other studies reported in fish-bearing streams could be due to a 35 number of factors. First, higher terrestrial input into headwater streams is expected given that headwaters have a smaller bank-full width on average than fish-bearing streams. Therefore, on average, headwaters have a greater percentage of stream-bed that is covered by, and receives inputs from, riparian canopy. I found that terrestrial inputs into headwater streams were 10.42 mg/m2/d from broadleaf canopies and 9.59 mg/m2/d from coniferous canopies. Given that terrestrial invertebrates in this study were found to be larger than aquatic invertebrates, I would expect these streams to have proportionally greater relative densities of terrestrial invertebrates in the drift. Secondly, given selective fish predation on terrestrial invertebrates in fish-bearing reaches, the relative density of terrestrial invertebrates in fish-bearing streams should be lower than headwaters, where there is not selective fish predation. Given this evidence for higher drift densities of terrestrial invertebrates in headwater streams than fish-bearing streams, headwater inputs entering larger fish-bearing streams would be expected to increase terrestrial invertebrate densities in these downstream reaches. Fish responses to variation in headwater drift subsidies The export of invertebrate prey from Ashless headwaters to downstream reaches will have greater importance i f fish in those downstream reaches are generally food limited. Our study indicates that fish in coastal streams are indeed limited by food availability. The average ambient drift rate in each channel was 0.49 ±0.18 mg/m3, which is comparable to observed drift rates in other fish-bearing streams in British Columbia (Rosenfeld and Boss 2001). Fish receiving this amount of food (and no additional subsidies) experienced negative growth rates during summer (Figure 5). As the channels 36 were stocked at natural fish densities (Young et al. 1999), my results support previous findings that food is a critical environment factor limiting fish growth and viability. The growth response curve (Figure 5) produced in this study allows us to draw conclusions about potential fish responses to variation in invertebrate subsidies in comparison to other means of estimating fish response (Table 4). As expected, increased food resources led to increased fish growth. However, while resources were consumed in accordance with their abundance, it would also have been expected that fish receiving food rations at or above maximum daily ration at optimal temperature (6% biomass at 12°C) would have reached saturation capacity or maximum growth (Jobling 1994). The presence of excess food in these channels at the end of each day further supports this line of reasoning. Contrary to this expectation, fish grew at much lower rates than maximum in the channels. This finding suggests that other factors, such as capture efficiency or perhaps suboptimal habitat conditions in the channels, might interact with food availability to decrease expected growth rates (Hughes 1998). Bioenergetics modeling Sensitivity analysis identified the ratio of terrestrial to aquatic prey availability as the most sensitive parameter to consumer growth output response at high prey densities. Specifically, at high prey densities, altering the proportion of terrestrial and aquatic invertebrates in the streams caused the greatest percent change in consumer growth, above and beyond the influence of temperature. The reason for the magnitude of the consumer response is most likely due to the higher energetic quality of terrestrial invertebrates, given their greater caloric content (Cummins and Wuychuck 1971). Size 37 and visibility are also some of the determining factors in fish prey capture (Wilzbach et al. 1986, Angradi and Griffith 1989). This is demonstrated by the fact that fish communities will often shift to feeding in the benthos under conditions of limited visibility (Tippets and Moyle 1978), a strategy that is potentially less productive energetically (Nakano et al. 1999). When given the choice, fish show a preference for terrestrial invertebrates as prey (Hunt 1975, Allan et al. 2003) for a number of reasons. Not only are terrestrial invertebrates larger, but they also incur fewer foraging costs due to energy expenditure for prey capture and exposure time to predation. Further, unlike their aquatic counterparts, terrestrial invertebrates do not have developed mechanisms for predator avoidance in aquatic systems (Giller and Malmqvist 1998). Bioenergetics modeling of fish relative growth rates using measured stream parameters yielded a growth curve lower than that measured in our experiment (Figure 9b). Specifically, the area between the two curves was 123% different. Modeled growth curves might be lower due to default assumptions made by the model, such as uniform fish age and swimming costs, which are not applicable to this experiment. Fish in the Mayfly Creek channels were able to find slow water current refugia and therefore reduce active metabolic costs. Furthermore, the degree to which actual growth of fish exceeded the predictions implies that other foraging strategies (such as benthic feeding) might have been used to make up for low food availability. This hypothesis is further supported by the fact that, at maximum daily ration (input rate = 0.96 g/channel/day), actual and predicted relative growth rates converge. In particular, this convergence indicates cost-benefit relationships of drifting and benthic foraging strategies. Whereas apparent 38 increased benefits from benthic foraging strategies at low prey drifting densities may enable fish to exhibit higher than expected growth rates, at high prey densities, there appears to be adequate food in the drift to meet growth rate predictions given expected swimming costs. Therefore, comparison of our growth curve with that produced by the bioenergetics model highlights consumption differences and potential benthic foraging strategies of the cutthroat trout in this study. Further experimentation over a greater range of food availability would be needed to confirm this mechanism. Estimating the contribution of headwater systems to downstream fish populations Fish living in high- and low-order reaches are often food limited (Mason 1976, Bilby and Bisson 1987, Boss and Richardson 2002, this study). In general, as resource-limited aquatic systems efficiently utilize and process all available sources of energy (Lindeman 1942), it is likely that any additional inputs of food resources to these streams will support either increased individual fish growth or greater densities of fish populations. In this study, the proportion of terrestrial invertebrate biomass per unit area of headwater stream was greater than that reported elsewhere in fish-bearing streams. To estimate the potential magnitude of this contribution to downstream fish communities in my study area, I calculated the relative growth rate of fish receiving headwater inputs and those receiving only higher order stream invertebrate drift densities. To do so, I used the regression equation from the growth response curve in Figure 5 (relative growth rate = 0.225 (invertebrate drift density) - 0.029). Using an average drift density of 1.33 mg/m3 from headwaters and 0.23 mg/m3 from downstream reaches (Rosenfeld and Boss 2001), I 39 found that there is the potential for a 12-fold increase in relative growth rate of an individual fish feeding in the downstream reach, if fish efficiently consume the additional resources represented by headwater subsidies. We can examine the importance of headwater subsidies to downstream fish populations from another perspective by estimating the number of fish potentially supported by increased terrestrial invertebrate subsidies from headwater reaches. Given that invertebrates have been shown to drift on average 50 m per day (Brittain and Eikeland, 1988, Gerking 1994), we derived estimates of terrestrial invertebrate headwater inputs from an average 50 m reach of headwater above the junction with a fish-bearing stream. Input estimates per season and canopy type were calculated per 50 m reach based on average wetted width (1.74 m) and invertebrate input rates per streams. These resources can be converted into fish wet mass, number of fish supported by 50 m of headwater or even relative growth rate of individual fish (Table 4). Regardless of the conversion mechanism used, downstream fish populations clearly benefit from these additional resource inputs. For example, 50 m of headwater inputs above a junction with a fish-bearing reach has the potential to support, on average, more than 750 young of the year in broadleaf canopies and almost 575 young of the year in coniferous ones. In July, this estimate increases to over 800 fish in broadleaf canopies and more than 750 in coniferous ones (see Wipfli 2005). The combination of food limitation downstream and the abundance of upstream headwaters in a watershed make it highly likely that fish utilize resources flowing in from headwaters. This conclusion is supported by studies reporting the presence of higher 40 numbers of fish reported spending time in tributary junctions (Kiffney et al. 2006). These hotspots of biological activity presumably exist because they allow fish to maximize food intake by occupying nodes of productivity within the watershed. Further, activity hotspots of this type provide evidence that subsidies alter not only consumer growth but their abundance and distribution as well (Benda et al. 2004). Land Use Most current studies looking at invertebrate drift have focused on old growth coniferous forest or young alder and coniferous communities (Wipfli 2005, Wipfli and Musslewhite 2004). In this study, we investigated the link between these two successional stages - the second growth coniferous forest. Resource subsidies from second growth in BC appear to be less than those from either old growth or younger alder communities in Ashless streams in Alaska and similar to those in young growth coniferous forests in both Oregon and Alaska (Table 5). This finding suggests that invertebrate subsidies to stream fish populations might shift as forests age. Given enough time, long-term dynamics of stream recovery have been shown to restore invertebrate communities to their pre-harvesting state (Stone and Wallace 1998). Comparing our results to others at different successional stages, we therefore expect that as stream canopies shift from alder to coniferous, invertebrate productivity will decrease. As our second growth forests age, we would expect to see increased headwater resources. An assessment of resource availability from streams in the Pacific Northwest (Table 5) indicates that drift resources are, in general, extremely variable, making even regional generalizations difficult. In addition, sampling has been conducted in very different ways, and using different assessment techniques 41 ( A F D M vs. length/mass regressions). This study does not include winter invertebrate drift, which makes it difficult to develop true comparisons between different studies. Further study of different forest ages within the same geographic area would provide insight into this issue. Changes in riparian canopy type could result from future developments in forest harvesting practices, or even changes in climate. Frequency or severity of disturbance events might therefore have varying results on overall invertebrate subsidy. Conclusions The prevailing view is that of the connectivity of streams along the watershed network (Vannote et al. 1980, Power and Dietrich 2002). However, management of streams is typically on the local scale and prioritizes fish and fish habitat (Fausch et al. 2002). Given our understanding of natural linkages between upstream and downstream habitats (Gomi et al. 2002), long-term effects of logging are present not only directly in fish-bearing streams, but also in upstream reaches. This study has shown that upstream and downstream reaches are unidirectionally, but dynamically, linked in streams of the temperate rainforest of coastal British Columbia. In doing so, we question current approaches to management of fish habitat under the BC Forest and Range Practices Act. Given that there is a substantial flow of beneficial resources into downstream reaches from Ashless headwaters and that these headwaters typically receive more limited protections from forestry due to absence of fish, we propose that the management of fish habitat should extend upstream. Further, it appears that headwater streams are indeed providing an ecosystem service through support of fish production. 42 FIGURES AND TABLES FROM RESULTS Table 1. Comparison of physical stream characteristics in the two stream basin study areas. Values reported are means with ranges in brackets. M K R F = Malcolm Knapp Research Forest. MKRF Chilliwack Number of streams n = 21 n= 13 Latitude / Longitude 49°16'N, 122°34'W 49°10'N, 121°56'W Mean velocity (m/s) 0.13 (0.01-0.38) 0.27 (0.07-0.51) Mean gradient (°) 9.5 (0-21) 10(2-20) Mean temperature (°C) 7.81 (5.11-11.45) 7.17(6.10-8.26) Mean wetted width (m) 1.69 (0.82-2.76) 1.85 (0.68-3.46) Mean depth (m) 0.05 (0.02-0.26) 0.11 (0.05-0.20) Mean substrate size (m) a 0.10 (0.03-0.27) 0.06 (0.03-0.09) Largest particle based on measurements of b axis of five randomly encountered rocks per section of transect 43 Table 2. General linear mixed model output for comparisons of drift and terrestrial abundance and biomass. Comparisons were made by season, canopy type and invertebrate origin in the drift (aquatic or terrestrial). A l l second-order interactions were not significant. Month Canopy Origin Test Statistic P Test Statistic P Test Statistic P Drifting Invertebrates Abundance (no/m3) F,, 29.4 =9.75 0.004 Fl,28.5=0.89 0.354 Fi > 8 5 . 6= 15.13 0.0002 Biomass (mg/m ) F1,28.8=6.99 0.013 Fl , 28.8 =0.23 0.638 Fi,88=0.51 0.475 Terrestrial Invertebrates Abundance (no/m ) F i , 3 2 . i= 11.95 0.002 Fl , 32.7 =0.39 0.537 n/a n/a Biomass (mg/m2/day) Fi,3o=0.05 0.821 Fi.29., = 0.07 0.788 n/a n/a Table 3. Back-transformed least-squares means estimates of abundance and biomass based on season, canopy and origin of invertebrates in the drift. A l l estimates are based on general linear mixed models with restricted estimate maximum likelihood (REML) estimation. Month Canopy Origin Drift July May Broadleaf Coniferous Aquatic Terrestrial Abundance (no/m ) 2.01 1.29 1.73 1.49 1.52 1.14 Biomass (mg/m ) 1.54 1.16 1.37 1.30 1.22 1.18 Terrestrial Abundance (no/m2/day) 146.08 70.81 110.43 93.67 - -Biomass (mg/m2/day) 10.36 9.65 10.42 9.59 - -4^  Table 4. Different estimation methods of fish potentially supported by subsidies from an average 50 m reach of headwater stream3. Estimates are based on canopy type and season. Numbers reported are averages. Terrestrial Invertebrate Contributions May July Broadleaf Coniferous Broadleaf Coniferous Biomass estimation method Average Terrestrial Inputs (mg/m2/day) 22 16 23 21 Inputs per 50 m reach (mg/day) 1869 1400 1956 1852 Wet mass of trout (mg/m2/day)b 2071 1551 2168 2052 Y O Y fish (no. supported/day)0 766 574 802 759 1 year old (%biomass gain/day)d 1.60 1.20 1.68 1.59 a Average wetted width of streams is 1.74m b Estimate from Romaniszyn et al. 2006. Based on a 0.2 conversion efficiency of available food resource. c Estimate from Wipfli 2005. Based on consumption rates ranging from 5-20 mg/day. d Estimate from this study. Based on experimentally derived growth response curve. as Table 5. Abundance and biomass of invertebrate terrestrial input and drift from headwater streams in the Pacific Northwest. Terrestrial input measured in mg/m2/day. Drift measurements in mg/m 3 a. Location Canopy Terrestrial Aquatic Source Coastal old growth Southeast Alaska, USA X coniferous 37.00 1.70 Wipfli 1997, Wipfli 2005 Coastal young growth Allan et al. 2003, Piccolo and Wipfli Southeast Alaska, U S A 1 coniferous 83.30 1.00 2002 Allan et al. 2003, Piccolo and Wipfli Southeast Alaska, USA X Coastal deciduous 83.30 3.10 2002 Malcolm Knapp Research Forest, BC, Coastal young growth Canada* coniferous 9.59 1.30 This study Malcolm Knapp Research Forest, BC, Canada* Coastal deciduous 10.42 1.37 This study Coastal young growth Romero et al. 2005, N . Romero, Oregon Coast Range, USA coniferous 45.00 1.01 Oregon State University - pers. comm. Romero et al. 2005, N . Romero, Oregon Coast Range, USA Coastal deciduous 64.00 0.85 Oregon State University - pers. comm. Eastern B C , Canada Interior ** 5.10 McKay 2006 x Fishless stream a Collection methods varied from study to study. Some estimates are based on 24-hour drift samples while others are just from daytime samples. N A Legend "ft Chilliwack Study Area XT ' Malcolm Knapp Research Forest l Meters 0 6.25t2,500 25,000 37,500 50,000 Inset map shows location of study sites within British Columbia \ 1 i \ British Columbia ^ ^ ^ ^ ^ V \ Figure 1. Map of southwestern British Columbia with Malcolm Knapp Research Forest and the Chilliwack River Basin. Stars indicate the location of stream study areas. 48 Figure 2. (a) Abundance and (b) biomass of terrestrial invertebrate inputs per week throughout summer 2006 in two headwater streams of broadleaf and coniferous canopy Regression slopes are not significantly different from one another but are significantly different from zero. Week 1 = May 19-24, 2007, Week 15 = August 23 - September 1, 2007. 49 Figure 4. (a) Abundance and (b) biomass of terrestrial invertebrates inputs by season and canopy (least-square means ± 1SE). Bars with a '*' are significantly different at the 0.05 level. 51 Figure 5. Relative growth rates of cutthroat trout in Mayfly Creek channels based on experimental addition of terrestrial invertebrates. Regression line based on response per channel (n = 10) (see text for treatments included in the analysis). Individual fish (n = 30) plotted. Food addition at zero includes baseline drift in channels (0.038 g/channel/day on average). Fish in highest food addition (input rate = 0.96 g/channel/day) are receiving maximum daily ration. 52 18 O V) CD 0) t_ 0) TS 0) I-<D O) CO < 16 H 14 H 12 4 10 4 I , , iV tula j I i I l l i i l l ^llliiliil i — i — six two four Week Figure 6. Average daily temperature per two-week sampling period in Mayfly Creek experimental channels. Average temperature from each sampling period was significantly different from all other periods. Lower and upper edges of box represent first and third quartiles. Middle line in box represents median. Lines extending outside of box represent smallest and largest non-outlier observations. Dots outside box represent outliers. 53 1.0 -0.2 J-¥ , 1 1 , 1 1 , 1 0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6 Food Added (g/channel/day) Figure 7. Changes in relative growth rate across food treatments per time. Regression slopes are not significantly different from one another. Food addition at zero includes baseline drift in channels (0.038 g/channel/day on average). Fish in highest food addition (input rate = 0.96 g/channel/day) are receiving maximum daily ration. 54 Figure 8 (a) Prospective (conventional) and (b) retrospective (relative) sensitivity analysis output. At low prey densities, temperature and initial mass produce the largest changes in final mass. However, at high prey densities, changes in proportion of aquatic invertebrates produces the greatest change. Values plotted are percent change in final mass per food ration level over the 40 days of the simulation for each parameter. Estimates for prospective analysis are derived by calculating percent change in final mass from an initial model of average values to one where each parameter is increased individually by 10% of its average value. Estimates for retrospective analysis are derived by calculating percent change in final mass from an initial model of average values to one where each parameter is increased individually by 10% of its range. Terrestrial and aquatic invertebrate values were changed by increasing the proportion of each invertebrate category (while decreasing the proportion of the other invertebrate category). 55 0.5 H -0.5 H • Mayfly Creek O Bioenergetics Model s t e i o H -1.0 H B o.o 0.2 0.4 0.6 Food added (g/channel/day) 0.8 1.0 Figure 9 (a) Relative growth rates from Mayfly Creek experiment and Wisconsin Bioenergetics model. Values plotted are individual relative growth rates from each fish in the experiment. Food addition at zero includes baseline drift in channels (0.038 g/channel/day on average). Fish in highest food addition (input rate = 0.96 g/channel/day) are receiving maximum daily ration, (b) Relative growth rates from Mayfly Creek experiment and Wisconsin Bioenergetics Model. Values plotted are channel averages ± 1SE. Food addition at zero includes baseline drift in channels (0.038 g/channel/day on average). Fish in highest food addition treatment (input rate = 0.96 g/channel/day) are receiving maximum daily ration. 56 REFERENCES Allan, J. D. 1981. Determinants of diet of brook trout {Salvelinus fontinalis) in a mountain stream. Canadian Journal of Fisheries and Aquatic Sciences 38:184-192. Allan, J.D., M . S. Wipfli, J. P. Caouette, A . Prussian, and J. Rodgers. 2003. Influence of streamside vegetation on inputs of terrestrial invertebrates to salmonid food webs. Canadian Journal of Fisheries and Aquatic Sciences 60:309-320. Angelier, E. 2003. Ecology of streams and rivers. Science Publishers, Inc. Plymouth, UK. Angradi, T. R., and J. S. Griffith. 1990. Diel feeding chronology and diet selection of rainbow-trout {Oncorhynchus mykiss) in the Henrys Fork of the Snake River, Idaho. Canadian Journal of Fisheries and Aquatic Sciences 47:199-209. Bacon, P. J., W. S. C. Gurney, W. Jones, I. S. Mclaren, and A. F. Youngson. 2005. Seasonal growth patterns of wild juvenile fish: Partitioning variation among explanatory variables, based on individual growth trajectories of Atlantic salmon {Salmo salar) parr. Journal of Animal Ecology 74:1-11. Baxter, C. V. , K . D. Fausch, and W. C. Saunders. 2005. Tangled webs: Reciprocal flows of invertebrate prey link streams and riparian zones. Freshwater Biology 50:201-220. Benda, L. , N . L . Poff, D. Miller, T. Dunne, G. Reeves, G. Pess, and M . Pollock. 2004. The network dynamics hypothesis: How channel networks structure riverine habitats. Bioscience 54:413-427. Bilby, R. E., and P. A . Bisson. 1992. Allochthonous versus autochthonous organic-matter contributions to the trophic support of fish populations in clear-cut and old-growth forested streams. Canadian Journal of Fisheries and Aquatic Sciences 49:540-551. Bilby, R. E., and P. A . Bisson. 1987. Emigration and production of hatchery coho salmon {Oncorhynchus kisutch) stocked in streams draining an old-growth and a clear-cut watershed. Canadian Journal of Fisheries and Aquatic Sciences 44:1397-1407. Bilby, R. E., B. R. Fransen, P. A . Bisson. 1996. Incorporation of nitrogen and carbon from spawning coho salmon into the trophic system of small streams: Evidence from stable isotopes. Canadian Journal of Fisheries and Aquatic Sciences 53:164-173. Bisson, P. A . 1978. Diel food selection by 2 sizes of rainbow-trout {Salmo gairdneri) in an experimental stream. Journal of the Fisheries Research Board of Canada 35:971-975. Bland, R. G., M . J. Cuthbert, M . W. Cutkomp, W. G. Jaques, J. Lenieus, F. J. Stoner, J. Bamrich and E. T. Cawley. 1978. How to know the insects. Third Ed. McGraw-Hill. Boss, S. M . 1999. Summer resource limitation and over-winter movement and survival of stream-resident coastal cutthroat trout. MSc. thesis. Dept. of Forest Sciences. University of British Columbia, Vancouver, BC, Canada. 57 Boss, S. M . , and J. S. Richardson. 2002. Effects of food and cover on the growth, survival, and movement of cutthroat trout {Oncorhynchus clarki) in coastal streams. Canadian Journal of Fisheries and Aquatic Sciences 59:1044-1053. Brittain, J. E., and T. J. Eikeland. 1988. Invertebrate drift - A review. Hydrobiologia 166:77-93. Caraco, N . , and J. Cole. 2004. When terrestrial organic matter is sent down the river: The importance of allochthonous carbon inputs to the metabolism of lakes and rivers. Pages 301-316 in G. A . Polis, M . E. Power, and G. R. Huxel, editors. Food webs at the landscape level. University of Chicago Press, Chicago, Illinois, USA. Cartwright, M . A. , D. A . Beauchamp, and M . D. Bryant. 1998. Quantifying cutthroat trout {Oncorhynchus clarki) predation on sockeye salmon {Oncorhynchus nerka) fry using a bioenergetics approach. Canadian Journal of Fisheries and Aquatic Sciences 55:1285-1295. Caswell, H . 2000. Prospective and retrospective perturbation analysis: Their roles in conservation. Ecology 81:619-627. Christiansen, J. S., and M . Jobling. 1990. The behaviour and the relationship between food intake and growth of juvenile Arctic charr, Salvelinus alpinus L . , subjected to sustained exercise. Canadian Journal of Zoology 68:2185-2191. Cross, P. C , and S. R. Beissinger. 2001. Using logistic regression to analyze the sensitivity of P V A models: A comparison of methods based on African wild dog models. Conservation Biology 15:1335-1346. Cummins, K. W., and M . J. Klug. 1979. Feeding ecology of stream invertebrates. Annual Review of Ecology and Systematics 10:147-172. Cummins, K . W., and J. C. Wuycheck. 1971. Caloric equivalents for investigations in ecological energetics. Mitt. Int. Ver. Theor. Angew. Limnol. 18. 158 pp. Cummins, K. W., M . A . Wilzbach, D. M . Gates, J. B. Perry, and W. B. Taliferro. 1989. Shredder and riparian vegetation - Leaf litter that falls into streams influences communities of stream invertebrates. Bioscience 39:24-30. Cushing, C. E., and J. D. Allan. 2001. Streams: Their ecology and life. Academic Press. San Diego, California. Daily, G.C., Ed. 1997. Nature's services: Societal dependence on natural ecosystems. Island Press, Washington, DC. Dedual, M . , and K. J. Collier. 1995. Aspects of juvenile rainbow-trout {Oncorhynchus mykiss) diet in relation to food-supply during summer in the lower Tongariro River, New-Zealand. New Zealand Journal of Marine and Freshwater Research 29:381-391. 58 Edwards, E. D., and A. D. Huryn. 1996. Effect of riparian land use on contributions of terrestrial invertebrates to streams. Hydrobiologia 337:151-159. Egglishaw, H. J. 1964. The distributional relationship between the bottom fauna and Plant Detritus in Streams. The Journal of Animal Ecology 33:463-476. Engqvist, L . 2005. The mistreatment of covariate interaction terms in linear model analyses of behavioural and evolutionary ecology studies. Animal Behavior 70:967-971. Fausch, K. D., C. E. Torgersen, C. V . Baxter, and H. W. L i . 2002. Landscapes to riverscapes: Bridging the gap between research and conservation of stream fishes. Bioscience 52:483-498. Fish Bioenergetics 3.0. 1997. University of Wisconsin System Sea Grant Institute. Center for Limnology. Madison, WI, USA. Frady, C. H. 2005. Headwater stream macroinvertebrates of the H. J. Andrews experimental forest, Oregon. MSc. thesis. Dept. of Fisheries Science. Oregon State University, Corvallis, Oregon, USA. 97 pp. Frazer, G. W. 1999. Gap Light Analyzer Version 2.0. Simon Fraser University. Burnaby, British Columbia Canada. Gayraud, S., E. Herouin, and M . Phillippe. 2002. The clogging of stream beds: A review of mechanisms and consequences on habitats and macroinvertebrate communities. Bulletin Francais de la peche et de la pisciculture 365-66:339-355. Gerking, S. D. 1994. Feeding ecology of fish. Academic Press, Inc., California. Giller, P. S., and B. Malmqvist. 1998. The biology of streams and rivers. Oxford University Press Inc., New York. Gomi, T., R. C. Sidle, and J. S. Richardson. 2002. Understanding processes and downstream linkages of headwater systems. Bioscience 52:905-916. Grant, G. E., and A. L . Wolff. 1991. Long-term patterns of sediment transport after timber harvest, Western Cascade Mountains, Oregon, USA. Sediment and stream water quality in a changing environment: Trends and explanation (Proceedings of the Vienna symposium). IAHS Publ. no. 203. Gries, G., K. G. Whalen, F. Juanes, and D. L. Parrish. 1997. Nocturnal activity of juvenile Atlantic salmon (Salmo salar) in late summer: evidence of diel activity partitioning. Canadian Journal of Fisheries and Aquatic Sciences 54:1408-1413. Griffith, M . B., P. R. Kaufmann, A. T. Herlihy, and B. H. Hil l . 2001. Analysis of macroinvertebrate assemblages in relation to environmental gradients in Rocky Mountain streams. Ecological Applications 11:489-505. 59 Hansen, M . J., D. Boisclair, S. B. Brandt, S. W. Hewett, J. F. Kitchell, M . C. Lucas, and J. J. Ney. 1993. Applications of bioenergetics models to fish ecology and management: Where do we go from here? Transactions of the American Fisheries Society 122:1019-1030. Hartman, G. F., J. C. Scrivener, and M . J. Miles. 1996. Impacts of logging in Carnation Creek, a high-energy coastal stream in British Columbia, and their implication for restoring fish habitat. Canadian Journal of Fisheries and Aquatic Sciences 53:237-251. Harvey, B. C , J. L . White, and R. J. Nakamoto. 2005. Habitat-specific biomass, survival, and growth of rainbow trout (Oncorhynchus mykiss) during summer in a small coastal stream. Canadian Journal of Fisheries and Aquatic Sciences 62:650-658. Hassan, M . A. , M . Church, T. E. Lisle, F. Brardinoni, L. Benda, and G. E. Grant. 2005. Sediment transport and channel morphology of small, forested streams. Journal of the American Water Resources Association 41:853-876. Hawkins, C. P., M . L . Murphy, and N . H. Anderson. 1982. Effects of canopy, substrate composition, and gradient on the structure of macroinvertebrate communities in Cascade Range streams of Oregon. Ecology 63:1840-1856. Hogan, D. 1996. Channel assessment procedure field guidebook. Forest Practices Code of British Columbia co-published by BC Environment. Hughes, N . F. 1998. A model of habitat selection by drift-feeding stream salmonids at different scales. Ecology 79:281-294. Hunt, R. L. 1969. Overwinter survival of wild fingerling brook trout in Lawrence Creek, Wisconsin. Journal of the Fisheries Research Board of Canada. 26:1472-1483. Hunt, R. L. 1975. Food relations and behavior of salmonid fishes. 6.1. Use of terrestrial invertebrates as food by salmonids. Pages 137-151 in A. D. Hassler, editor. Coupling of land and water systems. Vol . 10. Springer-Verlag, New York. Jobling, M . 1994. Fish bioenergetics. St. Edmundsbury Press, Suffolk. Johansen, M . , J. M . Elliott, and A. Klemetsen. 2005. Relationships between juvenile salmon, Salmo salar L. , and invertebrate densities in the River Tana, Norway. Ecology of Freshwater Fish 14:331-343. Karlsson, O. M . , J. S. Richardson, and P. A. Kiffney. 2005. Modelling organic matter dynamics in headwater streams of south-western British Columbia, Canada. Ecological Modelling 183:463-476. Kiffney, P. M . , and J. S. Richardson. 2001. Interactions among nutrients, periphyton, and invertebrate and vertebrate (Ascaphus truei) grazers in experimental channels. Copeia 2001:422-429. 60 Kiffney, P. M . , C. M . Greene, J. E. Hall, and J. R. Davies. 2006. Tributary streams create spatial discontinuities in habitat, biological productivity, and diversity in mainstem rivers. Canadian Journal of Fisheries and Aquatic Sciences 63:2518-2530. Kline, T. C , J. J. Goering, and R. J. Piorkowski. 1997. The effect of salmon carcasses on Alaskan freshwaters. Pages 179-204 in A . M . Milner and M . W. Oswood, editors. Freshwaters of Alaska: Ecological Synthesis. Ecological Studies 119. Springer, New York. Knight, T. M . , M . W. McCoy, J. M . Chase, K. A. McCoy, and R. D. Holt. 2005. Trophic cascades across ecosystems. Nature 437:880-883. Liao, H. S., C. L. Pierce, and J. G. Larscheid. 2005. An empirical model for estimating annual consumption by freshwater fish populations. North American Journal of Fisheries Management 25:525-532. Lindeman, R. L. 1942. The trophic-dynamic aspect of ecology. Ecology 23:319-418. Littell, R. C. 2002. Analysis of unbalanced mixed model data: A case study comparison of A N O V A versus R E M L . Journal of Agricultural Biological and Environmental Statistics 7:472-490. Kutner, M . H., C. J. Nachtsheim, J. Neter, and W. L i . 2005. Applied linear statistical models. McGraw-Hill/Irwin, New York, N Y . MacDonald, L . H. 2000. Evaluating and managing cumulative effects: Process and constraints. Environmental management 26:299-315. Marczak, L . B. , R. M . Thompson, and J. S. Richardson. 2007. Meta-Analysis: Trophic level, habitat, and productivity shape the food web effects of resource subsidies. Ecology 88:140-148. Marden, M . , D. Rowan, and C. Phillips. 2006. Sediment sources and delivery following plantation harvesting in a weathered volcanic terrain, Coromandel Peninsula, North Island, New Zealand. Australian Journal of Soil Research 44:219-232. Mason, J. C. 1976. Response of underyearling coho salmon to supplemental feeding in a natural stream. Journal of Wildlife Management 40:775-788. Mason, C. F., and S. M . MacDonald. 1982. The input of terrestrial invertebrates from tree canopies to a stream. Freshwater Biology 12:305-311. MacKay, C. A . 2006. Effects of clearcut logging on organic matter and invertebrate drift exported from headwater streams in the interior of British Columbia. MSc. thesis. Dept. of Natural Resources and Environmental Studies. The University of Northern British Columbia, Prince George, BC, Canada. Merritt, R. W., and K. W. Cummins. 1996. An introduction to the aquatic insects of North America. Third Ed. Kendall Hunt, Dubuque, Iowa. 61 Moore, R. D. 2005. Small stream channels and their riparian zones in forested catchments of the Pacific Northwest: Introduction. Journal of the American Water Resources Association 41:759-761. Moore, R. D., D. L. Spittlehouse, and A. Story. 2005. Riparian microclimate and stream temperature response to forest harvesting: A review. Journal of the American Water Resources Association 41:813-834. Moore, R. D., and J. S. Richardson. 2003. Progress towards understanding the structure, function, and ecological significance of small stream channels and their riparian zones. Canadian Journal of Forest Research 33:1349-1351. Moyle, P. B., and J. J. Cech. 1988. Fishes: An introduction to ichthyology. Prentice Hall, Englewood Cliffs, New Jersey. Murphy, M . L. , and K. V . Koski. 1989. Input and depletion of woody debris in Alaska streams and implications for streamside management. North American Journal of Fisheries Management 9:427-436. Nakano, S., and M . Murakami. 2001. Reciprocal subsidies: Dynamic interdependence between terrestrial and aquatic food webs. Proceedings of the National Academy of Sciences of the United States of America 98:166-170. Nakano, S., Y . Kawaguchi, Y. Taniguchi, H. Miyasaka, Y . Shibata, H. Urabe, and N . Kuhara. 1999. Selective foraging on terrestrial invertebrates by rainbow trout in a forested headwater stream in northern Japan. Ecological Research 14:351-360. Palumbi, S. R. 2003. Ecological subsidies alter the structure of marine communities. Proceedings of the National Academy of Sciences of the United States of America 100:11927-11928. Piccolo, J. J., and M . S. Wipfli. 2002. Does red alder (Alnus rubra) in upland riparian forests elevate macroinvertebrate and detritus export from headwater streams to downstream habitats in southeastern Alaska? Canadian Journal of Fisheries and Aquatic Sciences 59:503-513. Polis, G. A. , and S. D. Hurd. 1996. Linking marine and terrestrial food webs: Allochthonous input from the ocean supports high secondary productivity on small islands and coastal land communities. American Naturalist 147:396-423. Polis, G. A. , W. B. Anderson, and R. D. Holt. 1997. Toward an integration of landscape and food web ecology: The dynamics of spatially subsidized food webs. Annual Review of Ecology and Systematics 28:289-316. Power, M . E., and W. E. Dietrich. 2002. Food webs in river networks. Ecological Research 17:451-471. 62 Price, K., A . Suski, J. McGarvie, B. Beasley, and J. S. Richardson. 2003. Communities of aquatic insects of old-growth and clearcut coastal headwater streams of varying flow persistence. Canadian Journal of Forest Research 33:1416-1432. Romaniszyn, E. D., J. J. Hutchens, and J. B. Wallace. 2007. Aquatic and terrestrial invertebrate drift in southern Appalachian Mountain streams: Implications for trout food resources. Freshwater Biology 52:1-11. Romero, N . , R. E. Gresswell, and J. L. L i . 2005. Changing patterns in coastal cutthroat trout (Oncorhynchus clarki clarki) diet and prey in a gradient of deciduous canopies. Canadian Journal of Fisheries and Aquatic Sciences 62:1797-1807. Rood, K., and R. E. Hamilton. 1995. Hydrology and water use for salmon streams in the Chilliwack/Lower Fraser Habitat Management Area, BC. Canadian Manuscript Report of Fisheries and Aquatic Sciences no. 2288. Rosenfeld, J. S., and S. Boss. 2001. Fitness consequences of habitat use for juvenile cutthroat trout: Energetic costs and benefits in pools and riffles. Canadian Journal of Fisheries and Aquatic Sciences 58:585-593. Smith, R. W., and J. S. Griffith. 1994. Survival of rainbow trout during their first winter in the Henrys Fork of the Snake River, Idaho. Transactions of the American Fisheries Society 123:747-756. Stone, M . K., and J. B. Wallace. 1998. Long-term recovery of a mountain stream from clearcut logging: the effects of forest succession on benthic invertebrate community structure. Freshwater Biology 39:151-169. Svendsen, C. R., T. Quinn, and D. Kolbe. 2004. Review of macroinvertebrate drift in lotic ecosystems. Wildlife Research Program, Environmental and Safety Division. Seattle City Light. Sweeney, B. W. 1993. Effects of streamside vegetation on macroinvertebrate communities of White Clay Creek in eastern North America. Proceedings of the Academy of Natural Sciences of Philadelphia 144:291-340. Sweeney, B. W., T. L. Bott, J. K. Jackson, L. A. Kaplan, J. D. Newbold, L. J. Standley, W. C. Hession, and R. J. Horwitz. 2004. Riparian deforestation, stream narrowing, and loss of stream ecosystem services. Proceedings of the National Academy of Sciences of the United States of America 101:14132-14137. Tippets, W. E., and P. B. Moyle. 1978. Epibenthic feeding by rainbow-trout (Salmo gairdneri) in Mccloud River, California. Journal of Animal Ecology 47:549-559. Vannote, R. L. , G. W. Minshall, K. W. Cummins, J. R. Sedell, and C. E. Cushing. 1980. River Continuum Concept. Canadian Journal of Fisheries and Aquatic Sciences 37:130-137. 63 Wallace, J. B., S. L . Eggert, J. L. Meyer, J. R. Webster. 1997. Multiple trophic levels of a forest stream linked to terrestrial litter inputs. Science 4:102-104. Ward, J. V. , and J. A . Stanford. 1982. Thermal responses in the evolutionary ecology of aquatic insects. Annual Review of Entomology 27:97-117. Ware, D. M . 1972. Predation by Rainbow-Trout (Salmo gairdneri) - Influence of hunger, prey density, and prey size. Journal of the Fisheries Research Board of Canada 29:1193-1201. Wilzbach, M . A. , K. W. Cummins, and J. D. Hall. 1986. Influence of habitat manipulations on interactions between cutthroat trout and invertebrate drift. Ecology 67:898-911. Wilzbach, M . A. , B. C. Harvey, J. L . White, and R. J. Nakamoto. 2005. Effects of riparian canopy opening and salmon carcass addition on the abundance and growth of resident salmonids. Canadian Journal of Fisheries and Aquatic Sciences 62:58-67. Winberg, G. G. 1956. Rate of metabolism and food requirements of fishes. Belorussian University, Minsk. Translated from Russian, 1960: Fisheries Research Board of Canada Translation Series 194, Ottawa. Wipfli, M . S. 2005. Trophic linkages between headwater forests and downstream fish habitats: Implications for forest and fish management. Landscape and Urban Planning 72:205-213. Wipfli, M . S. 1997. Terrestrial invertebrates as salmonid prey and nitrogen sources in streams: Contrasting old-growth and young-growth riparian forests in southeastern Alaska, USA. Canadian Journal of Fisheries and Aquatic Sciences 54:1259-1269. Wipfli, M . S., and J. Musslewhite. 2004. Density of red alder (Alnus rubra) in headwaters influences invertebrate and detritus subsidies to downstream fish habitats in Alaska. Hydrobiologia 520:153-163. Wipfli, M . S., and D. P. Gregovich. 2002. Export of invertebrates and detritus from Ashless headwater streams in southeastern Alaska: Implications for downstream salmonid production. Freshwater Biology 47:957-969. Wipfli , M . S., J. S. Richardson, and R. J. Naiman. 2007. Ecological linkages between headwaters and downstream ecosystems: Transport of organic matter, invertebrates, and wood down headwater channels. Journal of the American Water Resources Association 43:72-85. Wootten, R. J. 1990. Ecology of teleost fishes. Chapman and Hall, London. Young, K A , Hinch SG and TG Northcote. 1999. Status of resident coastal cutthroat trout and their habitat twenty-five years after riparian logging. North American Journal of Fisheries Management: 19:901-911. 64 Zhang, Y . X . , J. N . Negishi, J. S. Richardson, and R. Kolodziejczyk. 2003. Impacts of marine-derived nutrients on stream ecosystem functioning. Proceedings of the Royal Society of London Series B-Biological Sciences 270:2117-2123. Zimmerman, J. K. H., and B. Vondracek. 2006. Effects of stream enclosures on drifting invertebrates and fish growth. Journal of the North American Benthological Society 25:453-464. 65 CHAPTER 3: LOOKING FORWARD Implications Numerous studies (Wipfli et al. 2007, Cummins and Wilzbach 2005) have shown important ways in which to value headwater streams. That headwaters contribute a cool, clean water supply could become especially important in fish growth (Meeuwig et al. 2004) as climate change increases water temperatures in streams globally. Transport of woody debris and organic matter to downstream reaches creates important habitat structures (Gomi et al. 2002). Further, headwater streams have been shown to have the potential to increase biodiversity in downstream reaches (Meyer et al. 2007). In this study, we examined the strength of food resource linkages between headwater and downstream reaches, focusing on the cumulative effect of headwaters throughout a watershed (MacDonald 2000). In particular, our research provides evidence that headwater streams possess a tangible value in terms of supporting downstream resident cutthroat trout communities. As cutthroat trout have been identified as good indicator species for other salmonids (Guy 2004), drawing conclusions about their consumer response also permits us to make generalizations about overall salmonid growth and productivity in response to headwater subsidies. The study of subsidies allows us to consider on what scale we are evaluating ecosystems. Typically, S6 streams in British Columbia (<3m bank full width Ashless streams that are not in community watersheds), under the BC Forest and Range Practices Act, receive less riparian protection from harvesting practices than larger, fish-bearing streams. Similarly, in the United States, priority for protection lies with fish-bearing streams (Blinn and 66 Kilgore 2001). While the contribution of headwaters has been noted on many levels, this study provides a first measure at actual downstream response to these contributions, and the response appears to be substantial. In particular therefore, we must address the question of how riparian forests around Ashless headwater streams need to be managed. In terms of specifics, my study implies that introduction of broadleaf canopy post-logging might slightly augment invertebrate resources in the short-term. Successional dynamics however then appear to cause a decline in invertebrate resources, below initial levels. Given that food resources from broadleaf communities were also found to be potentially more variable, care should be taken in establishing stream community dynamics that might result in long-term fluctuating availability of resources. Given conservative invertebrate drifting distances, I recommend protecting headwater streams for 50-100 m reaches above fish-bearing junctions by managing them in the same way as the downstream reach. In order to mitigate adverse effects of logging on stream systems, current research points towards evidence that 10 m buffers of streams will preserve organic matter inputs (P. M . Kiffney and J. S. Richardson, University of British Columbia - pers. comm.) while at least 30 m of buffer is needed to provide protection against blowdown (Reid and Hilton 1998). A recent study, however, emphasized that sound and careful logging practices next to streams might be sufficient to prevent negative repercussions on instream communities (DeGroot et al. 2007). Although the ideal buffer width, as a function of the tradeoff between stream protection and harvesting, is not known, these studies are useful starting points. In buffering headwaters near downstream junctions, riparian canopies are not as severely modified and downstream reaches receive more stable food resources in the long-term. Further, sensitivity analysis shows larger consumer response at high prey densities. Therefore, prioritizing protection 67 of headwater streams when there are numerous headwaters in a stretch of fish-bearing stream should elicit the largest positive response in fish populations. Studying the impact of logging on headwater macroinvertebrate communities throughout the entire watershed network provides new insight into the kinds and magnitudes of ecosystem services offered by headwaters and the potential for cross-ecosystem subsidies to benefit salmonid populations. In particular, it allows us to fill in the story of headwater contributions by seeing the change in resources as a result of land use. Further, augmenting our knowledge of the extent to which fish populations make use of these subsidies is essential to increasing our understanding of fish energetics. Given the ecological and economic importance of fish to the Pacific Northwest, elucidating answers to these questions is key to future management and planning. Further, understanding the fundamental links between headwater and downstream habitats and different disturbance regimes brings new light onto the incorporation of spatial, temporal, and human dynamics into our current knowledge of natural processes. Suggestions for further research As with all research, answering one question often leads to the generation of more. While this study has taken a step in the direction of linking food resources from headwaters to consumers in downstream reaches, there are a number of ways in which this research can be carried forward. Specifically, the next logical step in this series of questions would be to conduct growth experiments on trout populations in non-experimental streams by manipulating invertebrate abundances in streams directly. In doing so, we would be able to draw more general conclusions about potential for trout response in the wild. On top of this, it would be interesting to study fish behavior around headwater junctions with 68 experimentally manipulated food resources. Specifically, one could test to see if fish spent proportionally more time around headwater stream junctions that had higher amounts of resources exported from them. Ideally, in this work, we wished to find a way to track or mark invertebrates in order to be able to look at gut contents of fish populations downstream and determine what proportion of their diet came from headwaters and from the fish-bearing stream. Should adequate marking material become available, direct contributions of headwaters to fish diets could be determined. Ultimately, the objective of this study was to look at the cumulative effects of invertebrate production on fish production. GIS work which could scale up invertebrate production while maintaining landscape microtopography (i.e. through a program like NetMap (Benda et al. 2007) could provide just such answers on cumulative impacts. Further, scaled up work could allow for commentary on both individual growth and density responses to subsidies. Study of population based growth rates as opposed to individual response can perhaps yield even more insight into determining the value of headwater subsidies. Investigating all of these questions would allow us to enhance our knowledge of upstream-downstream connections and the role of headwater streams in fish population production dynamics. 69 References Benda, L. , D. Miller, K. Andras, P. Bigelow, Q. Reeves, and D. Michael. 2007. NetMap: A new tool in support of watershed science and resource management. Forest Science 53:206-219. Blinn, C. R., and M . A . Kilgore. 2001. Riparian management practices - A summary of state guidelines. Journal of Forestry 99:11-17. Cummins, K. W., and M . A . Wilzbach. 2005. The inadequacy of the fish-bearing criterion for stream management. Aquatic Sciences 67:486-491. Degroot, J. D., S. G. Hinch, and J. S. Richardson. 2007. Effects of logging second-growth forests on headwater populations of coastal cutthroat trout: A 6-year, multistream, before-and-after field experiment. Transactions of the American Fisheries Society 136:211-226. Gomi, T., R. C. Sidle, and J. S. Richardson. 2002. Understanding processes and downstream linkages of headwater systems. Bioscience 52:905-916. Guy, T. 2004. Landscape-scale evaluation of genetic structure among barrier-isolated populations of coastal cutthroat trout (Oncorhynchus clarki clarki). MSc. thesis. Dept. of Fisheries Science. Oregon State University, Corvallis, Oregon, USA. MacDonald, L . H. 2000. Evaluating and managing cumulative effects: Process and constraints. Environmental management 26:299-315. Meeuwig, M . H. , J. B. Dunham, J. P. Hayes, and G. L. Vinyard. 2004. Effects of constant and cyclical thermal regimes on growth and feeding of juvenile cutthroat trout of variable sizes. Ecology of Freshwater Fish 13:208-216. Meyer, J. L. , D. L . Strayer, J. B. Wallace, S. L. Eggert, G. S. Helfman, N . E. Leonard. 2007. The contribution of headwater streams to biodiversity in river networks. 43:86-103. Reid, L. M . , and S. Hilton. 1998. Buffering the buffer. United States Department of Agriculture General Technical Report. PSW-GTR-168-Web. Wipfli, M . S., J. S. Richardson, and R. J. Naiman. 2007. Ecological linkages between headwaters and downstream ecosystems: Transport of organic matter, invertebrates, and wood down headwater channels. Journal of the American Water Resources Association 43:72-85. 70 APPENDICES Appendix A Appendix A. Prospective and retrospective sensitivity analysis input parameters. Average parameter values were used to create a baseline model for comparison with changed parameter values. Prospective sensitivity analysis used parameter values increased by 10% of average value. Retrospective sensitivity analysis relied on the range of each parameter, increasing the value by 10% of the range. This type of analysis therefore allows for incorporation of the variability of each parameter in the model and provides a more conservative means for comparing variable change. Parameter Average Average + 10% Range Range + 10% Temperature 12.39 13.62 7.21 13.11 Prey. Proportion Terrestrial Invertebrates 0.21 0.23 0.21 0.26 Aquatic Invertebrates 0.79 0.87 0.79 0.84 Initial mass (by % ration) 0.0 5.98 6.58 3.70 6.35 1.5 4.89 5.38 3.70 5.26 3.0 4.71 5.18 3.70 5.08 4.5 5.72 6.29 3.70 6.09 6.0 4.96 5.45 3.70 5.33 7.5 5.76 6.33 3.70 6.13 9.0 5.36 5.90 3.70 5.73 Prey Energy Densities (cal/g)* Aquatic Invertebrates 2657 Terrestrial Invertebrates 8401 Mysis shrimp 3418 * Average values extracted from Cummins and Wuycheck, 1971. —i Appendix B Appendix B. Prospective and retrospective sensitivity analysis output (Fish Bioenergetics 3.0, Madison, WI). Prospective sensitivity analysis used parameter values increased by 10% of average value. Retrospective sensitivity analysis relied on the range of each parameter, increasing the value by 10% of the range. This type of analysis therefore allows for incorporation of the variability of each parameter in the model and provides a more conservative means for comparing variable change. Model output was generated by sequentially modifying each parameter and simulating the model for forty days. Values reported are estimated final mass of individual fish using parameter values reported in Appendix A . Ration (%biomass/day) 0 1.5 3 4.5 6 7.5 9 Original Model 3.02 3.25 4.18 6.93 7.87 12.27 14.94 Prospective Temperature 2.72 2.91 3.75 6.23 7.08 11.07 13.49 Initial mass 3.36 3.61 4.65 7.69 8.74 13.61 16.57 Prey Proportion Terrestrial Invertebrates 3.02 3.28 4.26 7.12 8.16 12.82 15.74 Aquatic Invertebrates 3.02 3.13 3.90 6.25 6.88 10.39 12.28 Retrospective Temperature 2.88 3.09 3.98 6.61 7.50 11.71 14.28 Initial mass 3.23 3.52 4.55 7.42 8.52 13.13 16.06 Prey Proportion Terrestrial Invertebrates 3.02 3.32 4.38 7.42 8.61 13.70 17.03 Aquatic Invertebrates 3.02 3.17 3.99 6.47 7.19 10.98 13.10 t o Appendix C Appendix C. Prospective and retrospective sensitivity analysis3 of bioenergetics model parameters influencing predicted trout final masses (Fish Bioenergetics 3.0, Madison, WI). Prospective sensitivity analysis used parameter values increased by 10% of average value. Retrospective sensitivity analysis relied on the range of each parameter, increasing the value by 10% of the range. This type of analysis therefore allows for incorporation of the variability of each parameter in the model and provides a more conservative means for comparing variable change. Model output was generated by sequentially modifying each parameter and simulating the model for forty days. Values reported are percent change in growth due to a change in parameter value. Ration (%biomass/day) 0 1.5 3 4.5 6 7.5 9 Prospective Temperature -10.0 -10.3 -10.3 -10.0 -10.1 -9.8 -9.7 Initial mass 11.2 11.2 11.2 11.1 11.0 10.9 10.9 Prey Proportion Terrestrial Invertebrates 0.0 0.9 1.9 2.7 3.6 4.5 5.3 Aquatic Invertebrates 0.0 -3.4 -6.7 -9.7 -12.6 -15.3 -17.8 Retrospective Temperature -4.6 -4.8 -4.8 -4.6 -4.7 -4.5 -4.5 Initial mass 6.9 8.5 8.8 7.1 8.2 7.0 7.5 Prey Proportion Terrestrial Invertebrates 0.0 2.4 4.7 7.1 9.4 11.7 14.0 Aquatic Invertebrates 0.0 -2.3 -4.5 -6.6 -8.6 -10.5 -12.3 a Conventional sensitivity analysis increased each parameter by 10% of its mean value, while relative sensitivity analysis changed each parameter by 10% of its range 

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