Hydrobiologia (2012) 686:107–117 DOI 10.1007/s10750-012-1002-7
PRIMARY RESEARCH PAPER
The effects of riparian forest disturbance on stream temperature, sedimentation, and morphology Jered M. Studinski • Kyle J. Hartman Jonathan M. Niles • Patrick Keyser
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Received: 5 April 2011 / Revised: 22 December 2011 / Accepted: 15 January 2012 / Published online: 17 February 2012 Ó Springer Science+Business Media B.V. 2012
Abstract Forested headwater streams rely on their riparian areas for temperature regulation, woody debris inputs, and sediment retention. These products and services may be altered by disturbances such as timber harvest, windthrow, or development. This study investigated the effects of riparian forest disturbance by removing trees using 50 and 90% basal area harvests and by directly felling some trees into eight streams in eastern West Virginia. On summer afternoons, water temperature increased in the 50 and 90% BAH treatments at average rates of 0.18 and 0.79°C/ 100 m, respectively. The 90% BAH treatments had the potential to disrupt fish and invertebrate communities via increased water temperature. New roads and log landings associated with the riparian logging had no detectable effect on sedimentation or turbidity. Large
Handling editor: Nicholas R. Bond J. M. Studinski (&) K. J. Hartman Department of Forestry and Natural Resources, West Virginia University, PO Box 6125, Morgantown, WV 26506-6125, USA e-mail:
[email protected] J. M. Niles Department of Biology, Susquehanna University, 514 University Avenue, Selinsgrove, PA 17870, USA P. Keyser Department of Forestry, Wildlife, and Fisheries, University of Tennessee, 274 Ellington Plant Sciences Building, Knoxville, TN 37996, USA
woody debris (LWD) additions increased habitat complexity but no net increase in pool area was observed. Greater morphological instability was observed within the LWD addition sections as pools were both created and destroyed at significantly higher rates. Experimentally manipulating small riparian patches may be an analog for small-scale natural and anthropogenic disturbances. These common events are assumed to alter streams, but there are few experimental studies quantifying their effects. Keywords Riparian Disturbance Stream Morphology Temperature Sedimentation
Introduction Riparian zones exist at the interface between terrestrial and aquatic ecosystems. The boundaries of this interface are difficult to define, but the riparian zone can be thought of as a patchy area where the environmental variables are not as predictable or homogenous as its adjacent ecosystems (Naiman et al., 1988). This riparian interface regulates stream temperature (Hetrick et al., 1998; Moore et al., 2005), functions as a filter, buffer, and stabilizer (Keller & Swanson, 1979; Naiman & Decamps, 1997) and is the source of in-stream large woody debris (LWD). Forested headwater streams have particularly strong interactions with their riparian zones. The services and products of riparian areas can be altered by
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disturbances such as timber harvest (Davies & Nelson, 1994), development (Sponseller et al., 2008), windthrow (Grizzel & Wolff, 1998), fire (Minshall, 2003), and arboreal disease (Swanston, 1991). Disturbance plays a key role in stream ecology, and experimental studies are required to increase our understanding of these complex systems (Resh et al., 1988). This experiment focused on the effects of both riparian tree removal and instream LWD additions on stream temperature, sedimentation, and stream morphology. The reduction in riparian canopy cover results in higher stream temperatures, potentially altering fish and aquatic invertebrate (AI) communities (Hogg & Williams, 1996; Hawkins et al., 1997; Lessard & Hayes, 2003). Results from Sponseller et al. (2008) suggest disturbances as small as 200 m can affect stream biota. Cold-water taxa like plecopterans and brook trout (Salvelinus fontinalis) may be especially sensitive to increases in water temperature (Sweeney et al., 1986). In the eastern United States, brook trout populations are declining due to a suite of factors, including acidification, invasion of exotic species, and habitat loss (Larson & Moore, 1985; Haines & Baker, 1986). Increases in stream temperatures impose greater metabolic demands on brook trout (Hartman & Cox, 2008), which typically occupy low-productivity streams and often feed at or near maintenance ration (Sweka & Hartman, 2001a; Utz & Hartman, 2006). The riparian zone entraps and retains small particles, reducing sediment loads in streams (Rivenbark & Jackson, 2004, but see Keim & Schoenholtz, 1999). Riparian disturbance has been linked to increased sedimentation and turbidity (Jones et al., 1999; Swank et al., 2001), mainly due to the exposure of mineral soil when building of roads and log landings during forestry operations (Kochenderfer et al., 1997; Rivenbark & Jackson, 2004). Increased sediment loads can cause shifts in AI communities (Wood & Armitage, 1997), resulting in fewer Ephemeroptera, Plecoptera, and Trichoptera (EPT) taxa (Davies & Nelson, 1994; Kaller & Hartman, 2004). Brook trout have been shown to suffer from decreased growth as turbidity increases (Sweka & Hartman, 2001b). Additionally, trout reproductive success has been positively correlated with streams with low levels of fine sediments (Petty et al., 2005; Hartman & Hakala, 2006). Riparian disturbances can affect LWD inputs into streams. LWD improves detrital retention and
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increases habitat complexity, providing substrate for invertebrates and cover for fish (Bilby & Likens, 1980; Bilby, 1981; Hilderbrand et al., 1997). LWD is also a potential pool-forming mechanism (Bilby, 1984; Trotter, 1990; Gurnell et al., 1995), increasing coldwater refugia that may be critical for over-summering brook trout. Older forests donate more LWD to streams than younger forests (Silsbee & Larson, 1983; Flebbe & Dolloff, 1995; Richmond & Fausch, 1995), suggesting that any riparian disturbance that prevents succession to an old growth forest creates long-term LWD deficits in streams. Other disturbances like windthrows and arboreal diseases increase in-stream LWD and may improve stream quality for fish and invertebrates (Swanston, 1991), especially at smaller spatial scales where stream temperature does not substantially increase. Understanding the effects of riparian disturbance is especially useful for reducing anthropogenic impacts. Within the United States, each state develops their own regulations regarding anthropogenic riparian disturbances (Blinn & Kilgore, 2001; Lee et al., 2004), but the requirements and guidelines are generally similar in the establishment of a streamside management zone (SMZ). The SMZ is a riparian buffer within which disturbance and harvest are often restricted. For logging operations, West Virginia Division of Forestry (WVDF) best management practices (BMPs) require a road-free SMZ that extends at least 30 m from the stream bank for perennial and intermittent streams (WVDF, 2005). Within the SMZ, harvesting is not limited, but the operation of large equipment is not permitted. It is permissible to cut and remove trees from the SMZ. West Virginia BMPs require the removal of any LWD that may enter a stream during logging operations. However, the WVDF recognizes that LWD is important in stream ecology (West Virginia Division of Forestry, 2005). The objectives of this study were to quantify the effects of riparian tree removal at 50 and 90% basal area harvest (BAH) on stream temperature, sedimentation, and stream morphology. Furthermore, the effects of LWD additions on sedimentation and stream morphology were investigated. We expected the 90% BAH sections to be warmer and have smaller substrate than either the reference or 50% BAH sections. We also expected an increase in pool area in sections receiving the LWD additions. This experiment is valuable to land managers and ecologists who wish to
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better understand the effects of riparian disturbances in similar streams and adds to the limited information available regarding disturbance along streams in central Appalachia.
Methods Study area This study was conducted on eight small streams, all of which were tributaries of the Middle Fork River in Randolph County, West Virginia, United States. These first- and second-order streams are located on the Allegheny plateau at elevations of 705–868 m and have riffle-pool morphologies. The streams have slopes of 2.2–5.5%, drain 268–747 ha, have mean bank-full widths of 3.2–5.4 m, and have median substrate sizes of 6.7–12.1 cm. The area receives an average of 117 cm of precipitation annually. All streams occasionally experienced periods of low flow or no flow during the summer. The streams are surrounded by a mixed mesophytic forest (Van Sambeek et al., 2003) that was owned and managed for sawtimber and fiber production by MeadWestvaco Corporation but was sold to PennVirginia Corporation during the study. The second-growth riparian forest was 70–85 years old, and was dominated by black birch (Betula lenta), sugar maple (Acer saccharum), yellow birch (Betula alleghaniensis), and yellow poplar (Liriodendron tulipifera). The forest had a basal area of approximately 27.5 m2/ha (120 ft2/acre, A. Plaugher, personal communication). All experimental streams flowed over poorly buffered Pottsville geology. At the head of each stream, the West Virginia Department of Natural Resources historically added limestone sand to increase buffering capacity. All streams contained resident brook trout populations. The eight streams were divided into 250-m lower treatment, upper treatment, and reference sections
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(Fig. 1). A 100-m stream segment with an untreated riparian area separated the sections. Streams were randomly assigned treatments of either 50% or 90% BAH within the riparian zone. To adhere to West Virginia BMPs, the riparian zone was defined as being within 30 m of the stream edge regardless of slope or other factors. To assess the effects of LWD additions on stream morphology and sedimentation, either the upper or lower treatment section of each stream was randomly assigned the addition of LWD. The assignment was constrained so that half of the sections with 50 and 90% BAH had LWD added to the upstream treatment. During the harvest, loggers followed West Virginia BMP’s in all sections except those requiring LWD additions, where they haphazardly felled whole trees into the stream at approximately eight meter intervals. Riparian logging and LWD additions occurred August–December 2006, except on Kittle Creek, where treatments occurred in March 2007. Canopy cover and temperature Canopy cover was measured in order to relate streamsection-specific temperature changes to their respective coverage. Canopy data were collected in the summers of 2007 and 2008 (1 and 2 years postharvest) using a digital camera (Hewlett Packard model Q2190A, 85° maximum field of view). At approximate 15-m intervals, the canopy was photographed with the camera pointed directly upward, held 1 m above the stream, and adjusted to the widest fieldof-view. Using a custom black and white conversion in Photoshop (CS3 Extended version 10.0, Adobe Systems Incorporated), photographs were converted into binary images with branches and foliage black and sky and clouds white. The percent black pixels were recorded, yielding canopy coverage. Temperature data were collected during summer 2008 (2 years post-harvest). A HOBOÒ Water Temp
Fig. 1 Logging treatment design along each of eight Appalachian headwater streams. Four streams received 50% basal area harvest (BAH) and four 90% BAH. Each stream received woody debris additions in one of the treatment sections
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Pro data logger was fastened inside a section of white plastic pipe and anchored in a pool at the downstream and upstream edge of both treatment sections, and 250 m upstream from the top of the upper treatment section. Due to fiscal constraints the project was limited to 20 data loggers that had to be deployed in two events: four streams received data loggers on June 17, 2008, and four on August 7, 2008. Trends from the two sampling periods were combined for analyses and figures, as mean air temperatures differed by only 0.3°C. Temperature was recorded hourly for 32 days at each site. Total suspended solids and sedimentation Suspended solids were assessed through water samples collected from streams in spring and summer of 2007 and 2008 during five sampling events with covered a range of stream flows, including intense summer rain events and dry, base flow conditions. Very turbid water was observed during high-flow sampling events. At each stream, one liter of water was collected from the downstream edge of each section, and all sections were sampled within minutes of each other. The samples were then filtered through 0.7-lm pore pre-dried and pre-weighed glass fiber filters using a hand pump. The filters were dried to a constant weight and weighed to the nearest 0.0001 g. Sedimentation was quantified using substrate measurements in the summers of 2007 and 2008. Within each section, substrate diameter measurements were taken from 10 evenly spaced points along 10 transects throughout the section. Transects were perpendicular to stream flow. LWD and pool formation LWD surveys were conducted in summers of 2005 (pre-harvest), 2007, and 2008 (both post-harvest) to assess the impacts of the treatments. All LWD over 1 m in length, 10 cm in diameter and within the stream’s bank-full boundary was placed into one of seven size categories (Richmond & Fausch, 1995). Stream morphology was surveyed in summers of 2005, 2007, and 2008 during low flow conditions. Each habitat unit (i.e., pool, riffle, and run) had three width measurements, three depth measurements along each width measurement, and a length measurement
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(Dolloff et al., 1997). The maximum depth of pools was also recorded. To compare the rates of change in stream morphology, a simple instability index was created. This consisted of a proportional linear measure of change in habitat classification, calculated for each section. For example, if within a 250-m experimental section a 4.3-m pool was filled in and became a riffle, and a 5.7-m pool was created from another riffle, that section would have 10.0 m of instability, or proportionally, 0.04. The instability index is essentially a quantification of pools created and destroyed. Statistical analyses To assess changes in canopy coverage, data from 2007 to 2008 were combined to form an average canopy cover for each section. A repeated measures analysis of variance (ANOVA) with post hoc Wilcoxon signedrank tests was used to detect differences in canopy coverage between the treatments. Mean daily high stream temperature was calculated for each data logger location. Due to the loss and failure of some dataloggers and the resultant low sample sizes, testing for statistical differences between treatments was not practical. Regression analysis was used to model a relationship between canopy cover and temperature increase. Although spatial autocorrelation is present in the data, low sample sizes limited our analysis of the temperature data. Repeated measures ANOVA was used to compare the total suspended solids in each treatment section (harvest, harvest with LWD addition) to its corresponding reference section. To avoid very high turbidity values from masking trends at lower turbidities, each sampling event for each stream was considered an independent sample. For each section, the median substrate size was calculated and compared with its reference. Data 2007 and 2008 were combined by section. Repeated measures ANOVA was used to compare the treatment sections with their corresponding references. LWD volume per section was calculated and compared over time and between treatments using two-way repeated measures ANOVA with post hoc Wilcoxon signed-rank tests. Year and treatment were both fixed variables and were included to detect both initial differences between the treatments and any effects of the manipulations for 2 years following the
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treatments. Two-way repeated measures ANOVA was used to compare the average maximum pool depth of each section with its corresponding reference before and after the riparian harvest. Pool area was calculated and compared over time and between treatments using two-way repeated measures ANOVA. Similar to the LWD analysis, year and treatment were both fixed variables. Relative instability was calculated by comparing the 2005 and 2008 stream morphology surveys. Treatments were compared with their references using repeated measures ANOVA with post hoc Wilcoxon signed-rank tests. Although ecological data are rarely normally distributed, ANOVAs were used for all repeated measures tests. However, to more conservatively interpret the results, all post hoc analyses were performed with nonparametric tests. All statistical analyses were performed in the R programming language (R Development Core Team, 2009). Spatial autocorrelation was difficult to avoid with this study, but whenever possible, repeated measures tests were used to compare treatment sections with their corresponding reference sections. The lack of independence between the two harvested sections of each stream was another statistical concern. The possible effects of this are explored in the discussion section.
Results Canopy cover and temperature The reference section, 50% and 90% BAH treatments produced 3 distinct groups of canopy coverage (P \ 0.001, Fig. 2); each treatment was significantly different from the others (P \ 0.001–0.005). The variation in canopy coverage for the 90% BAH treatments was likely due to the lack of harvest in a few steep, inaccessible riparian areas. Canopy cover played a role in stream warming and temperature fluctuations. When stream temperatures were at their daily maxima, water temperature change appeared negligible as the stream flowed through the uncut areas upstream of the treatments (Fig. 3). Comparisons between stream sections were made during the hour in which the downstream section had the warmest water. Daily maximum temperatures increased as water flowed through both the 50 and 90% BAH treatments (0.18 and 0.79°C/100 m,
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Fig. 2 Basal area harvest (BAH) treatment effects on canopy coverage. Symbols represent mean values with 95% confidence intervals. Each treatment type was significantly different from the others (post hoc pairwise comparisons P \ 0.001–0.005)
respectively). After passing through the upstream treatment sections and warming, stream temperature decreased in the 100-m unharvested sections (average of -0.28°C/100 m). After passing through both 250-m treatment sections and the 100-m unharvested section, the maximum daily stream temperatures had increased an average of 0.5 and 3.7°C in the 50 and 90% BAH treatments, respectively (Fig. 3). Stream averages of daily maximum temperatures at the bottom of the downstream treatment sections varied from 15.9 to 18.0°C in the 50% BAH treatments and 17.2–20.4°C in the 90% BAH treatments. High stream temperatures were observed more often in the 90% BAH sections than in the 50% BAH sections (Fig. 4). During the warmest days, the 90% BAH sections occasionally reached temperatures [24.0°C, and at times increased [8.0°C from the reference section to the bottom of the furthestdownstream treatment. A regression model indicated a positive relationship between stream temperature and canopy cover (P \ 0.001, r2 = 0.49, Fig. 5). Average summer stream temperatures exhibited the same trend as maximum daily temperatures, although they were less extreme due to the inclusion of nighttime data. Average stream temperatures changed by -0.02°C/100 m in the reference sections, 0.03°C/
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Fig. 3 Cumulative temperature change as water flowed through unlogged and logged sections of headwater streams. Differences were calculated from daily high temperatures over a one-month period during summer. The ‘‘n’’ denotes the number of stream sections from which the mean increases were calculated, and the bars indicate the 95% confidence intervals
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respectively). In the 2 years following the treatments, some in-stream LWD formed large debris dams, and in one case, rerouted a stream down a new channel for approximately 80 m. In the four instances where the LWD addition section was in the upstream treatment, only once did LWD also increase in the downstream 100-m unlogged section. This indicated there was little downstream LWD movement. Pool area did not change in response to the LWD treatments (treatment 9 year interaction P = 0.942). However, the instability index, which compared cumulative linear change (2005 versus 2008), indicated significant differences between the treatments (P = 0.021, Fig. 6). Post hoc tests indicated there was greater instability in the LWD addition sections than in both the reference and harvest-only sections (P = 0.016 and P = 0.017, respectively). The treatments did not have an effect on average maximum pool depth (treatment 9 year interaction P = 0.546).
Discussion 100 m in the 50% BAH sections, and by 0.28°C/ 100 m in the 90% BAH sections. Total suspended solids and sedimentation The treatments had no effect on stream turbidity and sedimentation within the study reach. Repeated measures ANOVA found no differences in total suspended solids (TSS) when comparing treatment sections to their corresponding reference sections (50% BAH P = 0.835, 90% BAH P = 0.258, n = 20 for each harvest intensity). Median substrate size for each of the treatments did not differ from their corresponding references (50% BAH P = 0.413, 90% BAH P = 0.184, n = 4 for each harvest intensity). LWD and pool formation We detected significant treatment effects with greatly increased LWD in the LWD addition sections (treatment 9 year interaction P = 0.015). Volume of LWD was greater in the LWD addition sections compared to the corresponding reference sections in 2007 and 2008 (P = 0.025 for both). LWD volumes were not significantly higher than their corresponding reference sections in the harvested only (no LWD addition) sections in 2007 or 2008 (P = 0.093, P = 0.401,
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Canopy cover and temperature The canopy reduction that resulted from the BAH treatments affected stream temperature, although the relationship was more variable as canopy cover decreased (Fig. 5). Much of the variability in the streams’ responses to canopy removal is probably due to differences in groundwater inputs, stream size, and aspect (LeBlanc et al., 1997; Moore et al., 2005). Had the treatments been extended beyond 250 m, a nonlinear relationship would likely have been observed (Moore et al., 2005), which limits the extrapolation of these data. Due to the way the data were collected, it was possible to track changes in water temperature within each section. Even though the treatment sections within a stream were not independent, increases in temperature were consistent and appeared additive. The observations in this study fall within the wide range of changes observed or predicted in western North America (Moore et al., 2005), Japan (Sugimoto et al., 1997), and Australia (Davies & Nelson, 1994), although direct comparisons between studies are difficult due to differences in treatment and sampling methodologies. The 50% BAH sections, which created an average 71% canopy cover, resulted in an average temperature
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Fig. 4 Frequencies of hourly temperature observations of reference and treatment sections (50 and 90% basal area harvest) in Appalachian headwater streams. Data were gathered during two sampling events in summer 2008
Fig. 5 A significant inverse relationship was detected between canopy cover and change in mean daily high stream temperature (P \ 0.001, r2 = 0.49). The regression line is y = -0.012x ? 1.070. Symbols represent the basal area harvest (BAH) treatment for each stream section
increase rate of 0.18°C/100 m during the warmest part of the day. This increase may be biologically inconsequential at the scale of this study (250 m treatments), but had the treatment been extended further, it is possible that a 50% BAH could increase stream temperatures to harmful levels. Interestingly, the 100-m untreated section sufficiently cooled the stream so as to negate any temperature increases from the 250-m upstream treatment section. This suggests that small, patchy disturbances where \50% basal area is removed may not have long-reaching effects on downstream temperatures. The 90% BAH sections, which created an average 35% canopy cover, resulted in an average temperature increase rate of 0.79°C/100 m during the warmest part of the day. This may have been severe enough to stress and displace stream biota. Brook trout prefer temperatures under 16°C, above which growth rates decrease
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In some cases, aquatic macroinvertebrate density may have a positive correlation with temperature (Kreutzweiser et al., 2005), but the relationship would be nonlinear, have a maximum threshold, and would be highly dependent on initial stream conditions. Periphyton biomass usually increases with increasing stream temperature (Rosemond, 1994; Kiffney et al., 2003), but again, this is dependent on initial stream conditions. Total suspended solids and sedimentation
Fig. 6 A repeated measures ANOVA, comparing treatments to reference sections in eight Appalachian headwater streams, indicated that pools were created and destroyed at higher rates in stream sections that received LWD additions (P = 0.021). Instability was calculated by summing the length of pools that were created or destroyed and then dividing by the total length of the treatment section
(McCormick et al., 1972; Meehan & Bjornn, 1991). Mortality increases at temperatures above 18°C (Peterson et al., 1979), and the upper lethal limit for brook trout is approximately 25°C (Hynes, 1970; McCormick et al., 1972; Power, 1980). In the 90% BAH sections, stream temperatures regularly exceeded 20°C with temperatures as high as 24.5°C observed (Fig. 4). Increases in temperature, especially when coupled with acidic conditions and low productivity systems, could be harmful. Cold-water refugia would be important within the 90% BAH sections. Although the data loggers were placed in pools, it is possible that brook trout could find cooler refugia within the treatment sections (i.e., pools with greater groundwater influence), or move out of the sections. It is likely that portions of 90% BAH sections were at least temporarily reduced in habitat quality for brook trout due to high temperatures. The effects of increased temperature on aquatic macroinvertebrate communities and periphyton biomass are difficult to predict as there are often simultaneous effects from increased light and sediment. Sweeney et al. (1986) found that a consistent temperature of 15°C was fatal to some plecopterans and that decreased growth was observed when water was warmed 3°C over ambient stream temperature. Davies and Nelson (1994) found a reduction in plecopterans and some ephemeropterans as temperature increased.
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Following the experimental manipulations, the streams did not appear to have increased TSS or sedimentation, although these data should be interpreted with caution. Sedimentation and TSS data collection began 3–6 months after the disturbance (and subsequent reseeding of exposed soil) occurred. Additionally, the treatments in this study extended only 30 m from the stream, leaving the remainder of the watershed relatively undisturbed. Sedimentation from logging operations is the result of building roads and log landings (Kochenderfer et al., 1997; Kreutzweiser & Capell, 2001), which was minimal in this study because only a small portion of the watershed was recently disturbed. Thus, these data do not represent the immediate or cumulative impacts of watershedwide disturbances. It should be noted that this portion of the project was especially susceptible to the lack of independence between treatment sections (the upstream treatment may have affected the downstream treatment) but despite this, no increases in TSS or sedimentation were observed. LWD and pool formation While the LWD additions were substantial, there was no net gain in pool area or change in pool maximum depth. These results were similar to other LWD addition studies in high-gradient streams (Hilderbrand et al., 1997; Sweka & Hartman, 2006; Sweka et al., 2010). However, pools were created and destroyed at higher rates in the LWD treatment sections. Stream morphologies are generally at dynamic equilibrium, which is dependent upon factors such as flow, slope, and substrate size (Leopold et al., 1964; Chih, 1971). Hilderbrand et al. (1997) suggest that careful placement of LWD may increase pool area above a stream’s equilibrium, especially in lower gradient streams. However, the LWD treatments in this study were
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haphazard and essentially random. The habitat surveys suggest there was almost no downstream movement of LWD from one treatment section to another, indicating that upstream treatment sections had little effect on downstream treatment sections. LWD, even when not associated with pool creation, has a positive effect on fish and invertebrates (Flebbe & Dolloff, 1995; Gurnell et al., 1995; Rosenfeld et al., 2000; Hernanadez et al., 2005). This study’s copious LWD additions greatly increased habitat complexity within pools and riffles (personal observation) and reduced the long-term debris deficit that occurs in most streams following logging (Silsbee & Larson, 1983; Flebbe & Dolloff, 1995; Hedman et al., 1996). Much LWD was not recorded in our surveys because it was adjacent to or spanned the stream channel, but will serve as an important future source of LWD. However, with an in-stream half-life of approximately 20 years (Hyatt & Naiman, 2001), streams may still suffer from LWD deficits as harvest rotations in this region are typically [65 years. Due to its simplicity, haphazard LWD additions may be a simple way improve stream quality.
Conclusion The results of this study are applicable to small scale (\250 m) riparian disturbances along other forested, low-order streams with similar slopes, climate, and forest type. In riparian disturbances where canopy cover is reduced, the rapid increase in stream temperature is probably the most ecologically important effect. The relationship observed in this experiment between stream temperature and canopy cover may be useful in predicting the effects of small-scale logging operations, development, windthrow events, and arboreal disease outbreaks. Current West Virginia BMP’s, which do not limit riparian harvest, could be modified to further protect streams by creating harvest limits within the SMZ. Similar to other researchers investigating low-order streams, LWD additions did not result in a net gain in pool area but did increase habitat complexity. The felling of whole trees into streams may reduce long-term LWD deficits in streams flowing through managed forests and developed riparian areas. Further research extending this experimental design along spatial and temporal scales and extending the distance between the treatments
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would aid in understanding the effects of riparian disturbances on small streams. Acknowledgments I would like to thank MeadWestvaco, Penn Virginia Corporation, the US Forest Service, the National Fish and Wildlife Foundation, and the West Virginia DNR for financial support and access. I thank Dr. Todd Petty, Dr. Ray Hicks, Dr. John Strazanac, Holly Henderson, John Howell, Bryan Olejasz, Dr. George Merovich, Jason Stolarski, Charlie Russell, Garrett Staines, Andy Orsborn, Aaron Nemeyer, and Geoff Weichert for their guidance and hard work. I would also like to thank the anonymous reviewers for their suggestions and improvements to this manuscript.
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