2010 NORTHEASTERN NATURALIST 17(4):593–614 Differential Effects of Urbanization and Non-natives on Imperiled Stream Species Scott A. Stranko1,*, Susan E. Gresens2, Ronald J. Klauda1, Jay V. Kilian1, Patrick J. Ciccotto1, Matthew J. Ashton1, and Andrew J. Becker1 Abstract - The distribution of imperiled stream fish, crayfish, salamander, and freshwater mussel species of Maryland streams in relation to urban land cover and nonnative species was investigated. Over the last 30 years, extinction or extirpation of 13 stream animal species (including the endemic Etheostoma sellare [Maryland Darter]) was observed within the Piedmont region of Maryland, where urbanization has spread extensively outward from Baltimore and Washington, DC, and many non-native species have become established. The presence of imperiled species in this area was correlated with urbanization and non-native species occurrence. However, correlations with land-cover data were stronger than with non-native occurrence. The majority of sites with imperiled species contained less than 10% urban land cover and less than 5% impervious land cover in their catchments. In contrast, stream reaches with non-native species spanned the entire gradient of urban, agriculture, and forested land cover, with the majority of non-native species persisting in streams with over 60% urban and 40% impervious land cover. The persistence of rare species coincident with introduced species in rural portions of Maryland indicates that habitat degradation (like that typically accompanying urbanization) may be the most important threat limiting the distributions of the rarest species that remain in these streams. Limits on urbanization in areas with rare species are needed to maintain regional and global biological diversity. This is particularly important in areas like Maryland that are anticipating extensive human population and urban growth over the next 30 years. Introduction Rates of extinction and imperilment tend to be considerably higher for aquatic taxa than for terrestrial taxa (Allan and Flecker 1993, Palmer et al. 2000, Ricciardi et al. 1998). In Maryland, for example, 88% (14 of 16) of freshwater mussels and 41% (29 of 71) of native fish species are imperiled (included on the state list of rare, threatened, and endangered species; Maryland Department of Natural Resources 2007). Nearly 10% (6 of 71) of Maryland’s native freshwater fishes are presumed extirpated or extinct, including Maryland’s only endemic vertebrate, Etheostoma sellare (Radcliffe and Welsh) (Maryland Darter) (Helfman 2007, Jelks et al. 2008, Raesly 1991). Providing adequate protection for extant imperiled species is crucial to ensure the conservation of current stream biological diversity. Landscape-level diversity (beta diversity; i.e., the differences in community composition between different sites), is an important component 1Maryland Department of Natural Resources, 580 Taylor Avenue, C-2, Annapolis, MD 21401-2352. 2Department of Biological Sciences, Towson University, 8000 York Road, Towson, MD 21252. *Corresponding author - sstranko@dnr.state.md.us. 594 Northeastern Naturalist Vol. 17, No. 4 of regional biodiversity. Loss of this landscape-level diversity is being observed in many terrestrial and aquatic habitats, and is referred to as biotic homogenization (McKinney 2005; Rahel 2000, 2002; Taylor 2004). Human activities cause biotic homogenization by increasing both the dispersal of foreign species and local extirpation of native species. Although the impacts of biotic homogenization on richness (alpha diversity) of a local assemblage may be positive, negative, or neutral, the impacts serve to increase the similarity of faunas across the landscape. Environmental stress and humaninduced habitat degradation allow populations of tolerant native species to expand (Rahel 2002) and favor proliferation of non-native species (Baltz and Moyle 1993, Blair 2001, Byers 2002, Dukes and Mooney 2000, Dunham et al. 2003, Knutson et al. 1999, Limburg and Schmidt 1990, Morley and Karr 2002). Faunas can also become less similar if different species are introduced into separate locales, or species common to two regions are lost from one region and not the other. There are many potential mechanisms for species losses from one region or another. Habitat, hydrologic, and chemical degradation that accompany urbanization and biotic interactions with non-native species have been major causes of aquatic species extirpation from regions of the United States. The decline of 91–94% of imperiled fish species in the United States has been attributed to habitat degradation, while nonnative fish interactions may have affected 53 to 70% (Lassuy 1995, Reed and Czech 2005, Wilcove et al. 1998). The high rate of imperilment among freshwater mussels (Ricciardi and Rasmussen 1999) has been linked to poor land-use practices, habitat and flow alteration, and invasive species (Bogan 1993, Brim Box and Mossa 1999, Ricciardi et al. 1998, Strayer 1999, Watters 2000). Poor land-use practices and habitat degradation have also contributed to declines in stream salamanders (Rohr et al. 2004, Southerland and Stranko 2008, Willson and Dorcas 2003) and crayfishes (Taylor et al. 1996, 2007), but invasive species may represent the most important threat to native crayfish diversity (Capelli 1982; Capelli and Munjal 1982; Holdich 1988; Perry et al. 2001, 2002; Taylor et al. 1996, 2007). For all stream-dwelling taxa, habitat degradation likely exacerbates the negative influence non-native species have on native species (Moyle and Williams 1990). However, the relative importance and differential effects of these stressors towards extirpations and imperilment of aquatic fauna, in a regional context, are currently not well understood. Despite the importance of imperiled (rare, threatened, and endangered) fauna to regional biodiversity, rare species are often excluded in examinations of broad landscape-scale alterations on stream quality due to the paucity of records, which hampers rigorous statistical analyses. The deletion of “rare” species is considered a legitimate “ecological transformation” (McCune and Grace 2002) to prepare data for multivariate analyses whose goal is data reduction, i.e., representation of taxa-rich community data in a smaller number of synthetic axes. In such cases, the questions of interest 2010 S.A. Stranko et al. 595 focus on how the composition of groups of species may respond to environmental factors, and thus rare species provide little information and add more “noise” to the community response. In contrast, hypotheses regarding richness and diversity emphasize rare species as much as common ones, and deletion of any species would be inappropriate. Although species richness metrics for a local habitat give equal weight to both rare and common species, regionally rare and imperiled species do not contribute any extra weight to total richness indices, even though they are much more important for conservation than common, widespread species. The dearth of data and distribution records for rare and endangered stream-dwelling species further contributes to the problem. Data from the Maryland Department of Natural Resources’ Maryland Biological Stream Survey (MBSS) provide a unique opportunity to address this issue, with a large number of records for many imperiled and introduced fish, amphibian, crayfish, and mussel species that can be used to examine the impacts of landscape alteration on these taxa. The purpose of this paper is to use the MBSS data set to describe differential patterns of imperiled and non-native species distributions as they relate to land cover and non-native species in Maryland’s Atlantic slope. As a case study, we also examined patterns of land use and non-native species introductions within watersheds coincident with apparent Maryland Darter extinction. We hypothesize that extirpation/imperiled status of native species has a stronger correlation with urbanization than with the introduction of non-native species. In addition to testing this hypothesis, we document the degree to which landscape alteration is correlated with differences in rare and introduced species’ distributions in Maryland. Methods We used fish, crayfish, salamander, and freshwater mussel records from all 2740 sites selected via stratified random sampling of first- through fourthorder stream reaches by the MBSS during 1994–2007 east of the Appalachian Mountains. The Ohio River drainage portion of Maryland, west of the Appalachian Mountains, was not included because of major zoogeographical differences between the Ohio drainage to the west and Atlantic Slope drainages to the east (Hocutt and Wiley 1986). Data were collected using standard MBSS protocols. A detailed explanation of MBSS sampling protocols can be found in Stranko et al. (2007). In brief, backpack electrofishing, with two passes in each 75-m-long site, was used to collect fishes, crayfishes, and stream salamanders. Visual encounter surveys within the 75-m site for at least 15 minutes were used to collect freshwater mussels and to supplement the salamander and crayfish electrofishing catch. These data were used to compile current Maryland stream assemblages by site and physiographic region. The percent of forest, agriculture, and urban land-cover data from the 2001 National Land Cover Data-set (NLCD; Homer et al. 2007) were extracted for catchments upstream of each site, which were drawn by hand using digital USGS 7.5-minute topographic quadrangle maps. 596 Northeastern Naturalist Vol. 17, No. 4 Extensive literature reviews (Cooper 1983; Harris 1975; Jenkins and Burkhead 1994; Lee et al. 1976, 1980, 1981; Merideth and Schwartz 1960; Ortmann 1909, 1919; Schwartz et al. 1963) were used to reconstruct historical (before European settlement of North America) fish, crayfish, salamander, and mussel assemblages in Maryland’s physiographic provinces east of the Appalachian Mountains. Imperiled species are those on Maryland’s list of rare, threatened, and endangered animals (Maryland Department of Natural Resources 2007). The relative contribution of species extirpations versus introductions to biotic homogenization was estimated separately for three physiographic regions: Coastal Plain, Piedmont, and Highland (Fig. 1). Data from the Appalachian Plateau, Ridge and Valley, and Blue Ridge physiographic provinces were combined to make the “Highland” region based on Southerland et al. (2006), who found these three component provinces to be ecologically similar. We then estimated the change in similarity of species assemblages that has occurred among these regions due to species extirpations and introductions, using the approach of Rahel (2000). This method consisted of first comparing the current assemblages of the three regions using Jaccard’s similarity index. We then performed the comparison again without non-native species included and with species presumed to be extirpated added into the assemblage. We also compared the relative contribution of introductions and extirpations to the current similarity between regions by calculating Jaccard’s index with a) only extirpated species added or b) only non-native species removed. Spatial association of imperiled and introduced species with catchment land cover was determined for the Piedmont physiographic province using Figure 1. Map of the study area showing the locations of biological sampling sites and the three physiographic regions that were compared to examine potential biotic homogenization of streams, by region, in Maryland. 2010 S.A. Stranko et al. 597 data from 846 MBSS sites. We focused on the Piedmont because it was the only one of the three regions with a substantial amount (≥15%) of each major land-cover type (forest, agriculture, and urban). A total of 160 sites had a record of at least one imperiled species, whereas 581 sites had at least one non-native species. Because most of these sites included more than one type of land use, and land-use types may be spatially correlated on a local scale, we treated a catchment land-cover datum as a multivariate observation. Thus, our analyses sought to detect shifts in the suite of land-use types associated with the presence or absence of imperiled species. Multivariate analyses were conducted using PC-ORD 5.1 software (McCune and Mefford 2006). We used two approaches to analyze the circumstantial evidence provided by distributions of imperiled and introduced species. Initially, we defined two groups of watersheds, based on either the presence or absence of imperiled species. We tested the hypothesis of no difference in the suite of watershed land uses between these two predefined groups using a multi-response permutation procedure (MRPP), which is a non-parametric test of differences between two groups of multivariate observations (in this case, groups of stream sites). The significance of the test is determined by repeated random permutations of the data to yield a distribution of the test statistic under the conditions of the null hypothesis. The advantages of the MRPP test is that it does not require normally distributed data, nor homogeneity of variance within groups (McCune and Mefford 2006). We used an ordination technique, non-metric multidimensional scaling (NMS), to provide a graphical representation of catchment land cover for each stream, and to determine what shifts in land-use categories were responsible for the results of the MRPP test comparing streams with and without imperiled species. In our second approach, we compared associations between imperiled species and introduced species directly, using a contingency test (Zar 1999). The hypothesis of independent distributions of imperiled and introduced species across 846 Piedmont sites was tested using a “category 1” double dichotomy contingency test, which assumes a random sample of stream reaches in regards to presence of both imperiled and introduced species. We were concerned, however, that spatial autocorrelation in the distribution of these species within stream drainage networks and geographically in relation to urban areas could bias the interpretation of this test. Therefore, we also conducted a more conservative set of analyses using the Mantel’s statistic for correlation between two matrices (Mantel 1967). The Mantel’s tests were used to examine correlations in the distribution of these species groups, while accounting for possible non-independence of sites due to geographical proximity, drainage networks, or to spatial autocorrelation of habitat features. Thus, we tested three null hypotheses of no correlation: 1) between imperiled species and land use, 2) between non-native species and land use, and 3) between the occurrence of imperiled species and non-native species. 598 Northeastern Naturalist Vol. 17, No. 4 Randomization tests (based on 1000 random permutations) were used to establish the significance of each test. We used a case study to determine if landscape-alteration thresholds correlated with a presumed extinction were similar to thresholds correlated with the presences and absences of imperiled species in the MBSS dataset. This study involved the Maryland Darter and the two Maryland Piedmont watersheds where it was known to occur. These two watersheds are Deer Creek (37,700 ha) and Swan Creek (6820 ha). The Maryland Darter was found only in these two Maryland watersheds and is presumed extinct because the last record for the species was from 1988 (Raesly 1991). Land-cover data were available from the Maryland Department of Planning for three years spanning three decades: 1973, 1994, and 2000. Land-cover types in this data set were digitized from aerial photographs and satellite images and urban land use was verified using tax data from the appropriate time period. The minimum mapping unit was 4 ha, meaning that a unique land-cover type within a larger type must be at least 4 ha to be digitized separately. Presence or absence of the Maryland Darter during each of these years was estimated based on extensive surveys of historical habitats (Neely et al. 2003; Raesly 1991, 1992; US Fish and Wildlife Service 1985, 2007). Results Compared to historical estimates, total species and species within three of the four taxonomic categories increased in number or stayed the same for each physiographic region of Maryland (Appendix A). The only exception was salamanders, for which richness decreased or stayed the same (Table 1). Extirpations of species were highest (13) in the Piedmont and lowest (none) in Highland streams. Twenty non-native species were found in Highland, 19 in the Piedmont, and 13 in the Coastal Plain. Seventeen of the nineteen species introduced to the Piedmont were also introduced in the Highland. Introductions were dominated by fish species (19 of the 24 total non-native species). Introductions of non-native species had a consistent homogenization effect by increasing similarity among regions compared to historical estimates (Fig. 2). Introductions also contributed more than extirpations to changes in faunal similarity. The extirpation of different native species from the Piedmont and the Coastal Plain resulted in greater dissimilarity rather than homogenization. This decrease in similarity offset the five percent increase in similarity caused by the same species having been introduced into the two regions. The largest increase in overall taxonomic similarity (10%) occurred between the Piedmont and Highlands (from 67% to 77% similarity). The increase in similarity between the historically least similar regions (Highlands and Coastal Plain) was also relatively large (37% to 44%). Sixteen percent of sites with virtually no urban land cover (less than 1%; n = 266), contained at least one imperiled species (Fig. 3). Imperiled species were found at only three sites with more than 20% urban land cover (n = 133) and at no sites with more than 25% urban land cover (n = 126). The majority 2010 S.A. Stranko et al. 599 Table 1. Number of fish, mussel, crayfish, and salamander species in Maryland's Atlantic drainage by physiographic region. Historical refers to species present prior to European settlement. Number of species Region Taxa Historical Current Introduced Extirpated Highland Fishes 41 58 17 0 Mussels 7 8 1 0 Crayfishes 3 5 2 0 Salamanders 7 7 0 0 Total 58 78 20 0 Piedmont Fishes 54 63 16 7 Mussels 13 11 1 3 Crayfishes 5 7 2 0 Salamanders 7 4 0 3 Total 78 84 19 13 Coastal Plain Fishes 53 61 10 2 Mussels 14 14 0 0 Crayfishes 5 7 2 0 Salamanders 5 4 0 1 Total 77 86 13 3 All regions Fishes 74 88 17 3 Mussels 16 17 1 0 Crayfishes 8 12 4 0 Salamanders 10 8 0 2 Total 108 125 22 5 Figure 2. Change in the total proportion of similarity (Jaccard’s) from historical to current biological assemblages in streams between three Maryland regions: Coastal Plain (cp), Piedmont (pied), and Highlands (high), as well as the proportion of similarity change attributable to extirpations and introductions. 600 Northeastern Naturalist Vol. 17, No. 4 of imperiled species (8 of 12) were found only at sites with less than 10% urban land cover. Consistent with the lack of correlation between land cover and non-native species, non-natives were found at about half (56%) of the sites with more than 20% urban land cover. Nearly three quarters (14 of 19) non-native Piedmont species were still found in heavily urbanized (≥60% urban cover; n = 43 sites) Piedmont streams. Nearly half (48%) of the sites with imperiled species also contained at least one non-native species. The MRPP test showed significantly different catchment land-cover types for streams inhabited by imperiled species vs. streams where imperiled species were absent (P < 0.0001). The affect size (A = 0.014) indicates that the magnitude of land-use difference was small. If the two groups of streams were internally homogeneous in relation to land use—and thus quite distinctive— the value of A would approach 1. In case of a random pattern of land use across groups, A would be near zero, and could take lower values in the case of extreme heterogeneity across groups (McCune and Grace 2002). Figure 3. Cumulative proportion of stream sites in five land-use categories with imperiled (a) and non-native (b) species represented with bars, along with numbers of species shown with black diamonds. Sample sizes within each land-cover category are: less than 1% = 266, less than 5% = 586, less than 10% = 666, less than 20% = 714; >20% = 133. 2010 S.A. Stranko et al. 601 The NMS ordination clearly arranged the stream sites according to the degree of catchment covered by forest, agricultural, and urban land use (Figs. 4, 5). Imperiled species were only collected from stream sites with low (less than 25% land cover) urbanization (Fig. 4), whereas sites with non-native species spanned the entire gradients of urban, agriculture, and forested land cover (Fig. 5). The two ordination axes, representing major patterns of variation in land cover, were interpreted using correlations of the original land-cover categories with the stream score on a given axis. Axis 1 depicts streams along a gradient from mostly agricultural (correlation coefficient r = -0.91; low scores) to mostly urban land use (r = 0.90; high scores). Axis 2 portrays streams along a second gradient from agriculture (r = -0.50) to forested land cover (r = 0.97). Overlay of symbols for the presence of imperiled species (Fig. 4) on the ordination scores emphasizes that imperiled species were collected only from streams with low catchment urbanization. When spatial autocorrelation of streams was not taken into account, a contingency test of independence of the co-occurrence of imperiled and introduced species indicated that the distribution, i.e., presence or absence, of Figure 4. Nonmetric multidimensional scaling (NMS) of urban, agriculture, and forested land cover from the National Land Cover Dataset (NLCD). Filled circles represent stream sites with imperiled species. 602 Northeastern Naturalist Vol. 17, No. 4 imperiled and introduced species were independent of each other (with Yates correction χ2= 0.634; 0.25 < α < 0.50). In contrast, the Mantel’s test, which does account for autocorrelation among sites, gave a significant, but low, negative correlation between the presence of imperiled and introduced species (r = -0.07, P < .001). Consistent with the results of the MRPP test on land-use differences, the Mantel’s test gave a significant positive correlation of the presence of imperiled species and the suite of catchment land-cover types (r = 0.14, P < 0.001). Land cover was not significantly correlated with the presence of non-native species according to Mantel’s test (r = 0.02, P = 0.13). The last record for the Maryland Darter in the Deer Creek watershed was from 1988 and the last record from the Swan Creek watershed was from 1965. Land-use data from the Maryland Department of Planning for 1973, 1994, and 2000 indicate that urbanization in the Deer Creek and Swan Creek watersheds increased over time, and replaced forested and agricultural land uses (Table 2). Deer Creek land use from 1973 provided the only land-use data for the period when the Maryland Darter was still known to occur in one of the two watersheds (Deer Creek). This was also the only year for either watershed with land-use data showing the urbanization to have been less than 10%. As Figure 5. Nonmetric multidimensional scaling (NMS) of urban, agriculture, and forested land cover from the National Land Cover Dataset (NLCD). Filled circles represent stream sites with non-native species. 2010 S.A. Stranko et al. 603 described above, most imperiled species are currently limited to catchments with less than 10% urbanization. While urban land use increased over the period when the Maryland Darter disappeared, all of the non-native species that we found in the Deer Creek (n = 58 sites) and Swan Creek (n = 11 sites) watersheds during 1994–2007 had been initially introduced over 50 years (Lee et al. 1976) prior to the species’ presumed extinction. Discussion Non-native species, urbanization, and many other factors contribute to the decline and loss of native species. Our findings illustrate, however, that many rare, threatened, and endangered species persist in portions of Maryland together with non-native species, but none occur in urban areas. By comparing statistical analyses with and without accounting for spatial autocorrelation, we acknowledge the presence of multiple confounding environmental factors. Thus, we emphasize that we are fully aware that one cannot prove anything using distributional data; however, one can build a case with circumstantial evidence—part of the “strength of evidence” approach used in stressor identification by the Environmental Protection Agency (Cormier et al. 2002, 2003). We do not intend to definitively conclude that the cause for the spatial and temporal patterns of species distribution and extirpation that we report can be solely attributed to urbanization, or that non-native species did not contribute to these patterns. However, our findings do lend support to those of other studies that document the drastic effects of even low levels of urbanization on stream species diversity (Angermeier et al. 2004, Klein 1979, Lucchetti and Feurstenburg 1993, Marchetti et al. 2006, Wang et al. 2001, Weaver and Garman 1994). Although many studies have documented negative correlations between non-native species and rare species (e.g., Lassuy 1995, Miller et al. 1989, Moyle 1976, Tyus and Saunders 2000, Wilcove et al. 1998), we found the correlation to be weak (r = 0.07). Furthermore, non-natives occurred commonly (48% of sites) with imperiled species at sites with little urban land cover. However, since non-native species are present throughout urban and non-urban watersheds, we were not Table 2. Percent urban, agriculture, and forest land cover for three different years in two watersheds where the Maryland Darter was historically known to occur. Land cover (%) Watershed Year Urban Agriculture Forest Deer Creek 1973A 4 62 34 (Last seen in 1994 11 43 46 1988) 2000 12 56 32 Swan Creek 1973 18 41 41 (Last seen in 1994 41 1 58 1965) 2000 41 1 58 AYear when the Maryland Darter was presumed to still live in the Deer Creek watershed, where the last collection was made in 1988. 604 Northeastern Naturalist Vol. 17, No. 4 able to investigate the affect of urbanization, in the absence of non-natives. We acknowledge that the combined effects of habitat degradation and competition with non-native species may have cumulative effects on many native species (Moyle and Williams 1990). Despite the fact that agriculture also alters watersheds, there was no obvious pattern with either imperiled or non-native species occurrences and the percentages of agriculture or forested land cover in this study. This result could be, in part, because possible legacy effects from past agriculture land use (Brush 2009) potentially eliminated stream species that could not tolerate extensive sedimentation from agricultural run-off that occurred before distributional data were available (Harding et al. 1998). Conversely, urban land cover has only recently begun spreading away from the major metropolitan centers of Baltimore and Washington, DC. Thus, we may be witnessing the extirpations of sensitive species as they face, for the first time, the increased flood frequency, erosion, and inputs of potentially toxic chemicals and sediment associated with urban run-off (Walsh et al. 2005). Having mostly the same non-native species in the highly urbanized Piedmont and the primarily rural Highland region, adds to the weight of evidence supporting the concept that urbanization may have contributed to many recent extirpations, although introduced species and other factors likely also influenced this distributional pattern. Species introductions have resulted in greater species richness in Maryland’s Atlantic drainage compared to assemblages present before European settlement. This finding is consistent with other studies (Gido and Brown 1999, Hobbs and Mooney 1998, Sax and Gaines 2003) showing increasing diversity with species introductions. However, in Maryland’s Piedmont, the number of apparent extirpations was substantial and greater than in the Coastal Plain or Highland regions. Some studies have reported greater differentiation of regional faunas, rather than homogenization, resulting from extirpations and/or introduction of different species in different regions (McKinney 2005, Taylor 2004). In this study, the contrary effects of extirpations and introductions combined together resulted in no net difference in similarity of Piedmont and Coastal Plain faunas compared to their historical assemblages. Nevertheless, there were local extirpations of particular species from each region, indicating that a lack of overall change in similarity, by itself, does not represent sustained biotic integrity at the local scale. Regardless of whether homogenization or differentiation occurs, local extirpations are the most obvious evidence of biodiversity loss, and extinctions are irreplaceable. The apparent extinction of the Maryland Darter is an example of permanent global biodiversity loss. This species exhibited traits common to many other severely imperiled stream fishes: endemism, small geographic distribution, ecological specialization, preference for benthic habitats, and small body size (Angermeier 1995, Burkhead et al. 1997, Etnier 1997). While these ecological traits may have made the long-term persistence of the Maryland Darter questionable even in the absence of human disturbances, these 2010 S.A. Stranko et al. 605 same traits likely also made the Maryland Darter more sensitive to even minor human disturbances (Helfman 2007) resulting from incremental increases in urban development. The weight of evidence from this study, which documents the apparent loss of the Maryland Darter and disappearance of 13 stream species from the highly urbanized Piedmont region, indicate that maintaining Maryland’s current stream species diversity may require strict limitations on urban development to certain areas. Many studies concur that sensitive taxa can be eradicated from streams at even low levels of urbanization (e.g., <5% impervious land cover; Angermeier et al. 2004, Southerland et al. 2005, Yoder et al. 2000). Urbanization has been shown to be correlated with increases in stream temperature (Galli 1991, Klein 1979, Schueler 1994, Stranko et al. 2008, Wolman and Schick 1967) and sediment (Fox 1974, Swarts et al. 1978) as well as less stable habitat (Booth and Jackson 1997, May et al. 1997), compared to streams in undeveloped areas. There are many factors which could be responsible for the local eradication of a rare species. However, while biological indices provide measures of ecosystem response to a gradient of disturbance, these aggregated indices are not very effective at identifying which factors directly caused an impairment or the loss of a species (Allan 2004). This limitation makes it even more difficult to design remediation and restoration strategies. Indeed, we are not aware of any study that documents the improvement of an urban stream to pre-urban condition following an attempt at rehabilitation. The process of rehabilitating instream habitat may itself be perceived by the biota as a long-lasting disturbance (Tullos et al. 2009). Rehabilitation of urban streams by planting riparian buffers and re-engineering instream habitat often gives only short-term results, with modest improvements in biotic conditions, unless catchment-wide actions are taken to divert stormflow runoff from entering streams and to intercept chemical pollutants in runoff (Booth 2005). Stormwater detention ponds provide only imperfect solutions to this problem: the cost of a detention pond large enough to reduce both peak flow as well as duration of stormflow inputs is prohibitive in established urban areas (Booth and Jackson 1997). Given the cost and difficulty of catchmentwide projects to disconnect impervious surfaces and stream channels in developed areas, preservation of catchments harboring biologically sensitive and imperiled stream fauna should receive the highest priority. Although the disproportionate sensitivity of certain “intolerant” species to urbanization has been recognized, most published limits of stream biology to urbanization have been based on correlations with biological index scores (Klein 1979, Morgan and Cushman 2005, Schueler 1994, Wang et al. 2001). Loss of a single species may not result in a substantial change to an index that combines information about many species. For example, fish index of biotic integrity (IBI) scores (Southerland et al. 2006) from Deer Creek in 1996 and a tributary to Swan Creek in 1997 were rated as “Good” (≥4 on a scale of 1–5). In fact, the IBI score of 5.0 in the Swan Creek tributary, where the Maryland Darter had not been collected for over 30 years, was the highest 606 Northeastern Naturalist Vol. 17, No. 4 possible score for this multi-metric assemblage index. Regardless of how highly generalized biological index scores rate streams, the apparent extreme sensitivity of the regionally rarest species makes knowledge of their distributions and ecological requirements vitally important to wise land-use planning in order to maintain both regional and even global biological diversity. Acknowledgments We thank D. Boward, R. Hilderbrand, R. Morgan, and three anonymous reviewers for reviewing this document; P. Angermeier for reviewing an early draft and providing recommendations; and R. Hilderbrand for guidance and assistance with generating land-use data. We also extend our gratitude to the long list of diligent, hard-working MBSS field-sampling crew members, seasonal employees, interns, and volunteers who collected the majority of the data presented in this study. This study was funded by State Wildlife Grant funds provided to the state wildlife agencies by US Congress, and administered through the Maryland Department of Natural Resources’ Natural Heritage Program. Literature Cited Allan, J.D. 2004. Landscapes and riverscapes: The influence of land use on stream ecosystems. Annual Review of Ecology, Evolution, and Systematics 35:257–284. Allan, J.D., and A.S. Flecker. 1993. Biodiversity conservation in running waters: Identifying the major factors that threaten destruction of riverine species and ecosystems. Bioscience 43:32–43. Angermeier, P.L. 1995. Ecological attributes of extinction-prone species: Loss of freshwater fishes of Virginia. Conservation Biology 9:143–158. Angermeier, P.L., A.P. Wheeler, and A.E. Rosenberger. 2004. 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Scientific Name Authority Common Name H P CP Fishes Acantharchus pomotis Baird Mud Sunfish R Alosa aestivalis Mitchell Blueback Herring N N Alosa mediocris Mitchell Hickory Shad N N Alosa pseudoharengus Wilson Alewife N N Alosa sapidissima Wilson American Shad N N Ambloplites rupestris Rafinesque Rock Bass I I Ameiurus catus Linnaeus White Catfish N Ameiurus natalis Lesueur Yellow Bullhead N N N Ameiurus nebulosus Lesueur Brown Bullhead N N N Anguilla rostrata Lesueur American Eel N N N Aphredoderus sayanus Gilliams Pirate Perch N Campostoma anomalum Rafinesque Central Stoneroller N N Carassius auratus Linnaeus Goldfish I I I Catostomus commersoni Lacepède White Sucker N N N Centrarchus macropterus Lacepède Flier R Channa sp. Snakehead I I Clinostomus funduloides Girard Rosyside Dace N N N Cottus caeruleomentum Kinziger, Raesly, Blue Ridge Sculpin N N N and Neely Cottus girardi Robins Potomac Sculpin N N Cottus sp. Checkered Sculpin N Cyprinella analostana Girard Satinfin Shiner N N Cyprinella spiloptera Cope Spotfin Shiner N N Cyprinus carpio Linnaeus Common Carp I I I Enneacanthus chaetodon Baird Blackbanded Sunfish R Enneacanthus gloriosus Holbrook Bluespotted Sunfish N Enneacanthus obesus Girard Banded Sunfish N Erimyzon oblongus Mitchell Creek Chubsucker N N N Esox americanus Gmelin Redfin Pickerel N N Esox luscius Linnaeus Northern Pike N N Esox masquinongy Mitchell Muskellunge N N Esox niger Lesueur Chain Pickerel I N Etheostoma blennioides Rafinesque Greenside Darter N Etheostoma caeruleum Storer Rainbow Darter I Etheostoma flabellare Rafinesque Fantail Darter N Etheostoma olmstedi Storer Tessellated Darter N N N Etheostoma sellare Radcliffe & Maryland Darter X X Welsh Etheostoma vitreum Cope Glassy Darter X R Etheostoma zonale Cope Banded Darter I Ethestoma fusiforme Girard Swamp Darter R Exoglossum maxillingua Lesueur Cutlip Minnow N N N Gambusia holbrooki Girard Eastern Mosquitofish N N N Hybognathus regius Girard Eastern Silvery Minnow N N N Hypentilium nigricans Lesueur Northern Hogsucker N N Ictalurus punctatus Rafinesque Channel Catfish I I I Lampetra aepyptera Abbott Least Brook Lamprey N Lampetra appendix DeKay American Brook Lamprey X R Lepisosteus osseus Linnaeus Longnose Gar R 2010 S.A. Stranko et al. 613 Scientific Name Authority Common Name H P CP Lepomis auritus Linnaeus Redbreast Sunfish N N N Lepomis cyanellus Rafinesque Green Sunfish I I I Lepomis gibbosus Linnaeus Pumpkinseed N N N Lepomis gulosus Cuvier Warmouth R Lepomis macrochirus Rafinesque Bluegill I I I Lepomis megalotis Rafinesque Longear Sunfish I I Lepomis microlophus Günther Redear Sunfish N N Luxilus cornutus Mitchell Common Shiner N N N Margariscus margarita Cope Pearl Dace R X Micropterus dolomieu Lacepède Smallmouth Bass I I Micropterus salmoides Lacepède Largemouth Bass I I I Moxostoma erythrurum Rafinesque Golden Redhorse N N Moxostoma macrolepidotum Lesueur Shorthead Redhorse N N Nocomis micropogon Cope River Chub N N Notemigonus crysoleucas Mitchell Golden Shiner I I N Notropis amoenus Abbott Comely Shiner R R R Notropis bifrenatus Cope Bridle Shiner X X Notropis buccatus Cope Silverjaw Minnow N Notropis chalybaeus Cope Ironcolor Shiner R Notropis hudsonius Clinton Spottail Shiner N N N Notropis procne Cope Swallowtail Shiner N N Notropis rubellus Agassiz Rosyface Shiner N Noturus gyrinus Mitchell Tadpole Madtom N Noturus insignis Richardson Margined Madtom N N N Oncorhynchus clarkii Richardson Cutthroat Trout I Oncorhynchus mykiss Walbaum Rainbow Trout I I Perca flavescens Mitchell Yellow Perch N N N Percina bimaculata Haldeman Chesapeake Logperch R Percina notogramma Raney & Hubbs Stripeback Darter X R Percina peltata Stauffer Shield Darter N N Percopsis omiscomaycus Walbaum Trout-perch X Petromyzon marinus Linnaeus Sea Lamprey N N Pimephales notatus Rafinesque Bluntnose Minnow N N Pimephales promelas Rafinesque Fathead Minnow I I I Pomoxis annularis Rafinesque White Crappie N N N Pomoxis nigromaculatus Lesueur Black Crappie I I I Rhinichthys atratulus Hermann Eastern Blacknose Dace N N N Rhinichthys cataractae Valenciennes Longose Dace N N Salmo trutta Linnaeus Brown Trout I I I Salvenilus fontinalis Mitchell Brook Trout R R R Sander vitreus Mitchell Walleye N N Semotilus atromaculatus Mitchell Creek Chub N N N Semotilus corporalis Mitchell Fallfish N N N Umbra pygmaea DeKay Eastern Mudminnow N Mussels Alasmidonta heterodon I. Lea Dwarf Wedgemussel X N Alasmidonta undulata Say Triangle Floater N N N Alasmidonta varicosa Lamarck Brook Floater N N Anodonta implicata Say Alewife Floater N N Elliptio complanata Lightfoot Eastern Elliptio N N N Elliptio fisheriana I. Lea Northern Lance N Elliptio lanceolata I. Lea Yellow Lance X N Elliptio producta Conrad Atlantic Spike N N N Lampsilis cardium Rafinesque Plain Pocketbook I I 614 Northeastern Naturalist Vol. 17, No. 4 Scientific Name Authority Common Name H P CP Lampsilis cariosa Say Yellow Lampmussel N N N Lampsilis r. radiata Gmelin Eastern Lampmussel N N Lasmigona subviridis Conrad Green Floater N X Leptodea ochracea Say Tidewater Mucket N Ligumia nasuta Say Eastern Pondmussel N Pyganodon cataracta Say Eastern Floater N N N Strophitus undulatus Say Creeper N N N Utterbackia imbecillis Say Paper Pondshell N N Crayfishes Cambarus acuminatus Faxon Acuminate Crayfish R Cambarus b. bartonii Fabricius Common Crayfish N N N Cambarus diogenes Girard Devil Crawfish N N Cambarus dubius Faxon Upland Burrowing Crayfish N Fallicambarus fodiens Cottle Digger Crayfish N Orconectes limosus Rafinesque Spinycheek Crayfish N N Orconectes obscurus Hagen Allegheny Crayfish R R Orconectes rusticus Girard Rusty Crayfish I I Orconectes virilis Hagen Virile Crayfish I I Procambarus acutus Girard White River Crawfish N Procambarus clarkii Girard Red Swamp Crawfish I Procambarus zonangulus Hobbs & Hobbs Southern White River I Crawfish Salamanders Cryptobranchus a. Daudin Eastern Hellbender X alleganiensis Desmognathus fuscus Green Northern Dusky Salamander N N N Desmognathus monticola Dunn Seal Salamander N Desmognathus ochrophaeus Cope Allegheny Mountain Dusky N Salamander Eurycea bislineata Green Northern Two-lined N N N Salamander Eurycea l. longicauda Green Longtail Salamander N N Gyrinophilus p. Green Northern Spring Salamander N porphyriticus Pseudotriton m. montanus Baird Eastern Mud Salamander X N Pseudotriton r. ruber Latreille Northern Red Salamander N N N Siren lacertina Linnaeus Greater Siren X X I