All connectivity, percent of assessment area with buffer, average width of buffer, and buffer condition metrics data combined into a single landscape context score for each of the 129 files and 47 reference sites evaluated using CRAM.
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Figure 43. Hydrology metric CRAM scores compared to reference-site data.
All water source, hydroperiod, and hydrologic connectivity metrics data combined into a single hydrology score for each of the 129 files and 47 reference sites evaluated using CRAM.
Evaluation of California Compensatory Wetland Mitigation Final Report (Review Copy) 139
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Figure 44. Physical structure metric CRAM scores compared to reference-site data.
All physical patch richness and topographic complexity metrics data combined into a single physical structure score for each of the 129 files and 47 reference sites evaluated using CRAM.
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Figure 45. Biotic structure metric CRAM scores compared to reference-site data.
All organic matter accumulation, biotic patch richness, vertical biotic structure, interspersion and zonation, percent invasive plant species, and native plant species richness metrics data combined into a single biotic structure score for each of the 129 files and 47 reference sites evaluated using CRAM.
Evaluation of California Compensatory Wetland Mitigation Final Report (Review Copy) 140
Figure 46. Mean percentage scores for each CRAM metric for mitigation sites (N=204) and reference sites (N=47). 0 10 20 30 40 50 60 70 80 90 100
Connectivity
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Mean Mitigation Site Data (N=204)Mean Reference Data (N=47) 0 10 20 30 40 50 60 70 80 90 100
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Mean Mitigation Site Data (N=204)Mean Reference Data (N=47) Evaluation of California Compensatory Wetland Mitigation Final Report (Review Copy) 141
Figure 47. Mean percentage scores for each CRAM metric by state board region. (N=204 mitigation sites) 0 10 20 30 40 50 60 70 80 90 100
Region 1 (N=11)Region 2 (N=33)Region 3 (N=21)Region 4 (N=24)Region 5F (N=7)Region 5R (N=10)Region 5S (N=35)Region 6T (N=2)Region 6V (N=2)Region 7 (N=2)Region 8 (N=23)Region 9 (N=34)Reference Site Data (N=47) Connectivity
% of Assessment
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Region 1 (N=11)Region 2 (N=33)Region 3 (N=21)Region 4 (N=24)Region 5F (N=7)Region 5R (N=10)Region 1 (N=11)Region 2 (N=33)Region 3 (N=21)Region 4 (N=24)Region 5F (N=7)Region 5R (N=10)Region 5S (N=35)Region 6T (N=2)Region 6V (N=2)Region 7 (N=2)Region 8 (N=23)Region 9 (N=34)Reference Site Data (N=47) Connectivity
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Evaluation of California Compensatory Wetland Mitigation Final Report (Review Copy) 142
Wetland Class
Depressional
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Lacustrine
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Riverine High
Riverine Low
Seep/Spring
Vernal Pool
Overall CRAM Percentage Score
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Figure 48. Overall CRAM percentage scores by wetland class (N=204 mitigation sites).
The dotted line represents the mean, the solid line the median. The 10th, 25th, 75th, and 95th percentiles are displayed.
Evaluation of California Compensatory Wetland Mitigation Final Report (Review Copy) 143
Impacted Required Obtained
Sum of Acreage
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Figure 49. Overall acreage obtained compared to required and impacted (N=143 files).
Certification Year
1992 1993 1994 1995 1996 1997 1998 1999 2000 2001 2002
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Figure 50. Acreage required and obtained by year (N=143 files).
Evaluation of California Compensatory Wetland Mitigation Final Report (Review Copy) 144
Certification Year
1992 1993 1994 1995 1996 1997 1998 1999 2000 2001 2002
Average Mitigation Ratio
0 2
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Figure 51. Average mitigation ratios of required and obtained acreage by certification year as determined from our detailed permit file review.
In 2002, one file was removed that had 0.035 acres of impact and 4.30 required and obtained acres, yielding an anomalous mitigation ratio of 122.9. The resulting sample size was N=142.
Region
1 2 3 4 5F 5R 5S 6T 6V 7 8 9
Sum of Acreage
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Figure 52. Acreage required and obtained by region.
Evaluation of California Compensatory Wetland Mitigation Final Report (Review Copy) 145
Impacted Required Obtained
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Figure 53. Total acreage impacted, required and obtained for 143 files assessed. Acreage also grouped by jurisdictional habitat classifications: “Waters of the US” and non-jurisdictional waters (“Non-Waters”).
Required acreage also consists of a “Non-Specified Riparian” component, which represents a mitigation requirement of riparian acres, but non-specified jurisdiction (waters or non-waters).
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Figure 54. Total acreage impacted and obtained, with jurisdictional habitats data for “Waters of the US” proportioned into wetland and non-wetland waters habitats, and data for “Non-Waters” proportioned into riparian and upland habitats.
N=138 files (There are five files for which wetland acreage was not specified for waters of the US).
Evaluation of California Compensatory Wetland Mitigation Final Report (Review Copy) 146
Permanent Temporary Created Enhanced Preserved
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Figure 55. Total acreage impacted proportioned into permanent and temporary impacts, and obtained acreage proportioned into created, enhanced and preserved, each proportioned further into Waters of the US and NonWaters of the US (N=143 files).
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Figure 56. Total acreage for Wetland and Non-Wetland Waters, each displaying impacted and obtained acreage. Impacted acreage is proportioned into permanent and temporary impacts, while obtained acreage is proportioned into created, enhanced and preserved.
N=138 files (There are five files for which wetland acreage was not specified for waters of the US).
Evaluation of California Compensatory Wetland Mitigation Final Report (Review Copy) 147
Regions
1 2 3 4 5F 5R 5S 6T 6V 7 8 9
Sum of Acreage
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Required Impact Waters Impact NW Obtained Waters Obtained NW N =
19 6
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26 15
13 15
Figure 57. Total acreage impacted and obtained proportioned into Waters of the US and Non-Waters of the US by state board region (N=143 files).
Total required acreage per region is also displayed. N displayed = number of files assessed per region for both impacted and obtained.
Evaluation of California Compensatory Wetland Mitigation Final Report (Review Copy) 148
Regions
1 2 3 4 5F 5R 5S 6T 6V 7 8 9
Sum of Acreage
0 20406080
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11 25
15 16
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25 2 2
3 12
15 Figure 58. Total acreage impacted and obtained proportioned into Wetland, Non-Wetland Waters, Riparian and Upland jurisdictional habitats by state board region. Total required acreage per region is also displayed. N displayed = number of files assessed per region for both impacted and obtained. Total N=138 files (There are five files for which wetland acreage was not specified for waters of the US).
Evaluation of California Compensatory Wetland Mitigation Final Report (Review Copy) 149
Percent of Acreage Met
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Figure 59. Correlation analysis between percentage of acreage requirement met and average 401 permit compliance score (N=123 files).
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Overall CRAM Score
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Figure 60. Correlation analysis between percentage of acreage requirement met and overall filewide CRAM score (N=128 files).
Evaluation of California Compensatory Wetland Mitigation Final Report (Review Copy) 150
Average 401 Compliance Score
0 20 40 60 80 100
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Figure 61. Correlation analysis between average 401 permit compliance score and overall filewide CRAM score (N= 110 files).
Percent of 401 Conditions Met
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Figure 62. Correlation analysis between percentage of 401 permit conditions met and overall filewide CRAM score (N=110 files).
Evaluation of California Compensatory Wetland Mitigation Final Report (Review Copy) 151
Average Mitigation Plan Score
0 20 40 60 80 100
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Figure 63. Correlation analysis between average mitigation plan compliance score and overall filewide CRAM score (N=77 files).
Evaluation of California Compensatory Wetland Mitigation Final Report (Review Copy) 152
Site Implementation Conditions
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Figure 64. Correlation analysis between overall filewide CRAM percentage score and average 401 permit compliance score for four of the permit condition categories.
Sample sizes per condition category are as follows: for site implementation N=57, site maintenance N=18, site protection N=25, success/performance standards N=42. See Methods for description of permit condition categories.
Evaluation of California Compensatory Wetland Mitigation Final Report (Review Copy) 153
Impacted CRAM Weighted Obtained
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0 50
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Figure 65. Total impacted acreage and obtained acreage weighted by condition score (N=129 files).
Wetland Type
Depressional
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Lagoon
Mean 401 Compliance Score
0 20
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N= 28 1 2 19 10 1
Figure 66. Mean 401 compliance score for different wetland types.
Includes invoked conditions; N=61 files
The University of Maine
DigitalCommons@UMaine
Publications
Senator George J. Mitchell Center for Sustainability
Solutions
7-2014
Creating Successful Vernal Pools: A Literature
Review and Advice for Practitioners
Aram J K Calhoun
University of Maine, [email protected]
J. Arrigoni
State University of New York College of Environmental Science and Malcolm L. Hunter Jr.
University of Maine, [email protected]
C. Richter
Eastern Kentucky University
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Repository Citation
Calhoun, Aram J K; Arrigoni, J.; Hunter Jr., Malcolm L.; and Richter, C., "Creating Successful Vernal Pools: A Literature Review and Advice for Practitioners" (2014). Publications. 16.
https://digitalcommons.library.umaine.edu/mitchellcenter_pubs/16 Creating Successful Vernal Pools: A Literature Review and Advice for Practitioners A. J. K. Calhoun, J. Arrigoni, R. P. Brooks, M. L. Hunter, S. C. Richter A. J. K. Calhoun, M. L. Hunter
Department of Wildlife, Fisheries, and Conservation Biology, University of Maine, Orono, ME 04469, USA e mail: [email protected] J. Arrigoni
Department of Environmental and Forest Biology, State University of New York College of Environmental Science and Forestry, Syracuse, NY 13210, USA
P. Brooks
Riparia, Department of Geography, Pennsylvania State University, University Park, PA 16802, USA C. Richter
Department of Biological Sciences and Division of Natural Areas, Eastern Kentucky University, Richmond, KY 40475, USA
1 Abstract
Vernal pools provide critical breeding habitat for amphibians adapted to temporary waters, but they seldom receive the same level of protection as permanent wetlands. In response to continued degradation and loss of pools, managers often attempt to mitigate losses through pool creation or restoration. However, mitigation efforts often fail to provide suitable aquatic habitat for vernal pool amphibians. We review the literature on pool creation in northeastern and central North America, highlighting how and why constructed pools often fail to support amphibian related objectives. We recommend that practitioners consider the complex ecology of pool ecosystems and the historical and current distribution of pools and other wetlands in their local context before designing pool mitigation projects. Using vernal pool creation as a mitigation option should be a last resort (i.e., when elimination of natural pools is unavoidable). Monitoring should be target specific and conducted for at least 5 years. Topographic, geologic, and other local factors affecting pool hydrology and ecology vary regionally; pool creation remains an imperfect science that will only advance by documenting failures and successes. We recommend an adaptive management approach to vernal pool creation in which the effectiveness of techniques is evaluated and refined based on research.
Keywords: Ambystomatid salamanders, Ephemeral wetlands, Pool breeding amphibians, Restoration, Seasonal depressions, Vernal pool creation, Wetland creation, Wood frogs 2 Introduction
Vernal pools in the central to northeastern regions of North America are seasonal wetlands in forested landscapes, free of permanent fish populations. As their name suggests, they typically are most fully inundated in the spring, and may also be referred to as temporary ponds, ephemeral wetlands, or various combinations thereof. Vernal pools provide the core breeding habitat for wood frogs (Lithobates sylvaticus), mole salamanders (Ambystoma spp.), and invertebrates (notably fairy shrimp [Anostraca]) adapted to life in temporary waters, but they are often inadequately protected because of their small size and ephemeral nature. When pools are lost or severely degraded, mitigation efforts often rely on pool creation or restoration to replace breeding habitat for pool associated amphibians. It is essential to distinguish creation of vernal pools from restoration because the context is often very different. Creation usually occurs either as a compensatory mitigation requirement accompanying the removal of an intact, naturally occurring vernal pool, or as part of a proactive program to augment or diversify habitat by building new pools. Vernal pool restoration projects, however, attempt to return a pool from an altered or degraded condition to a preexisting condition (National Research Council 1992; Brooks and Gebo 2013). This paper focuses on creation of vernal pools, and while creation and restoration approaches can vary dramatically, much of the relevant literature is also applicable to restoration.
While replacing a hectare of natural wetland with a hectare of created wetland may result in no net loss of wetland area, there is likely to be a net loss of ecological functioning (Hoeltje and Cole 2007; Dahl 2011; Moreno Mateos et al. 2012). In particular, early reviews by Kusler and Kentula (1990) and Kentula et al. (1992) demonstrated that wetlands created for compensatory mitigation often failed to reliably replace natural wetland functions. This shortcoming is addressed in more recent guidance from the U.S. Environmental Protection Agency (2008) regarding the national “Mitigation Rule.” Under most circumstances, U.S. federal and some state wetland regulations require that both area and functional impacts to vernal pools be mitigated. However, recent legal cases have raised the issue about whether hydrologically isolated wetlands should receive the same degree 3 of protection as those with surface connections to navigable waters (Mahaney and Klemens 2008). This situation threatens vernal pools in particular (e.g., Leibowitz 2003) and highlights the need for clear, effective mitigation strategies.
By definition, wetland restoration seeks to reestablish a wetland’s physicochemical, hydrological, and ecological functions (National Research Council 2001), and presumably vernal pool creation also seeks to establish new pools with functions similar to those of natural pools. However, vernal pools are among the most difficult wetland ecosystems to create or restore primarily because of their hydrological properties: particularly, the seasonal water regime that is their defining feature. Furthermore, vernal pool creation or restoration must be coupled with adjacent high quality post breeding habitat for the biphasic amphibian species that depend on these pools for breeding, but spend most of their lives in adjacent habitats (Lichko and Calhoun 2003; Semlitsch 2008; Simon et al. 2009).
For practitioners seeking guidance on vernal pool mitigation strategies, published information can be confusing and conflicting. Some manuals for vernal pool creation suggest creating pools is easy (Biebighauser 2003, 2011), yet peer reviewed papers to date report generally poor or ambiguous results in terms of providing breeding habitat to vernal pool associated amphibians (Lichko and Calhoun 2003; Petranka et al. 2003; Vasconcelos and Calhoun 2004, 2006; Petranka et al. 2007; Korfel et al. 2009; Gamble and Mitsch 2009; Denton and Richter 2013). In this paper, we review the current literature on pool creation in central to northeastern North America (referred to below as the region) as pool creation is common in this area yet the relevant literature has not been synthesized. However, analogous research on the creation of small, often temporary, pools has been undertaken in other parts of the world, e.g., Australia (Brainwood and Burgin 2009), Brazil (da Silva et al. 2011), California (Black and Zedler 1996), Spain (Ruhi et al. 2012), and the United Kingdom (Williams et al. 2008). The purpose of this paper is to fill this gap by:
(1) providing a summary of the major findings in the current vernal pool creation literature in our region with a 4 focus on key indicator amphibians (wood frogs and ambystomatid salamanders); (2) defining an ecologically appropriate endpoint for successful creation; (3) providing general recommendations to practitioners gleaned from the literature; (However, it is not the goal of this review to provide a “how to” manual for vernal pool creation); and (4) identifying research gaps that should be a focus of pool creation studies.
Although we have focused on one region in North America, we believe our recommendations for approaches to management and future research are relevant, at least conceptually, to other ephemeral wetlands around the world.
Vernal pools and key pool breeding amphibian species
Vernal pools provide critical breeding and nursery habitat, where amphibian eggs and larvae develop and emerge as metamorphs to adjacent forests, for wood frogs and species of mole salamanders. Pools may provide additional breeding sites for other species, including (but not limited to) four toed salamanders (Hemidactylium scutatum), American toads (Anaxyrus americanus), eastern spadefoots (Scaphiopus holbrookii), upland chorus frogs (Pseudacris feriarum), and fairy shrimp (Anostraca) (Calhoun and deMaynadier 2008). In addition, adult amphibians may use pools as stepping stones while moving to summer and winter refugia (Semlitsch 1998; Baldwin et al. 2006a). Pools in the region are typically located in forested settings where shaded, cool conditions with abundant leaf litter and coarse woody material support detrital based food webs in the pools (Palik et al. 2006; Magnusson and Williams 2006). Wood frogs, mole salamanders, and certain other species of amphibians are well adapted to these relatively nutrient poor systems where they have limited competition and reduced depredation from fish, other amphibians, and invertebrates associated with more permanent waters and more open canopy pools (deMaynadier and Houlahan 2008; Denton and Richter 2013). 5 How do we measure success?
A mitigation project should have defined goals, or endpoints of success. From our perspective, successful pool creation is demonstrated by reproduction and metamorphosis by key indicator species such as wood frogs and ambystomatid salamanders for the long term (i.e., not just the 5 year monitoring period). In practice, regulators and managers often use a lower standard of “success,” e.g., simply the presence of any amphibian species or the presence of a given target species within the first 2 years after creation (Lichko and Calhoun 2003); however, these metrics may be misleading. Created pools may be colonized by dispersing individuals from other breeding pools, or, in the case of onsite mitigation, by returning members of the original pool breeding population , but their presence is not indicative of successful breeding (Vasconcelos and Calhoun 2006).
Furthermore, monitoring protocols generally require as little as 2 years of monitoring post creation (Lichko and Calhoun 2003). Our review of the literature, however, underscores the importance of monitoring periods for a minimum of 5 years to incorporate demographic and environmental variability such as the interaction between climate (temperature and precipitation variability) and hydro-period. For example, it may take more than 5 years for green frog (Lithobates clamitans) or bullfrogs (L. catesbeianus), voracious predators of wood frog and salamander egg masses and larvae, to establish in semi-permanent or permanent pools at densities that generate a detrimental effect on the classic pool breeders using those same pools (Vasconcelos and Calhoun 2004; Denton and Richter 2013).
Below we review key factors that affect pool success from the perspective of key pool breeding amphibians. We look at both pool specific factors and larger scale issues associated with pool landscape setting.
We address pool specific concerns and landscape scale issues separately for simplicity while recognizing that there are many interactions between the two.
6 Pool specific factors affecting creation success
Hydrology
The hydrology of any given pool will be dictated by hydro-geomorphic setting (i.e., the landscape position of a pool in the watershed that dictates the major inputs and outputs of water to the system), as well as more proximate factors such as canopy cover (i.e., evapotranspiration), pool surface area, volume, depth, and connection to groundwater (Cole and Brooks 2000). Periodic drying (usually in summer, occasionally in autumn and winter) excludes resident populations of predatory fish, reduces the diversity and abundance of invertebrate predators, and limits competition and depredation from other amphibian species (Drayer 2012; Denton and Richter 2013; Julian et al. 2013). Additionally, amphibian diseases may be more likely to persist in permanent wetlands than in those that dry (Gray et al. 2009; Richter et al. 2013a). For these reasons, hydro-period (i.e., the timing and duration of inundation), is particularly critical to vernal pool ecosystem functions (Zedler 2000); even variations in vernal pool filling and drying on the order of 2 to 3 weeks can alter pool breeding amphibian community composition and the success of any given species (Paton and Crouch 2002; Babbitt et al. 2003; Baldwin et al. 2006a; Good 2006; Seigel et al. 2006; Timm et al. 2007). For example, wood frog and spotted salamanders deposit eggs early in the spring, but wood frog embryos and larvae develop more rapidly than spotted salamanders, necessitating a substantially longer hydro-period to support successful reproduction for spotted salamanders (Fig. 1). In contrast, marbled salamanders (Ambystoma opacum) deposit eggs in dry basins in autumn. Eggs hatch after vernal pools recharge, and overwintered larvae metamorphose on a schedule similar to wood frogs.
A number of studies report significantly longer hydro-periods for created pools versus natural pools as well as differences in water chemistry owing to miscalculation of water table depths or compacted soils impeding drainage (diMauro and Hunter 2002; Gamble and Mitsch 2009; Korfel et al. 2009; Drayer 2012; Denton 7 and Richter 2013). Longer hydro-periods, particularly semi-permanent to permanent, may mean increased numbers of amphibian species and invertebrates that prey upon target vernal pool breeders. Notably, green frog and bullfrog larvae, which take 2 to 3 years to develop, and eastern newt (Notophthalmus viridescens) adults are efficient predators of wood frog and ambystomatid egg masses and larvae, and may be abundant in created semi-permanent to permanent ponds (Kross 2014). These species are also potential reservoirs of disease (Daszak et al. 2004; Gahl et al. 2012). Denton and Richter (2013) reported that in Kentucky wood frogs and marbled salamanders were almost exclusively found in natural, ephemeral wetlands, whereas large frogs (e. g., green frogs and bullfrogs) and eastern newts were only found breeding in permanent, constructed wetlands. In this system, newts and ranid larvae consumed all wood frog eggs (100–190 clutches) and many ambystomatid eggs in four created pools over a 4 week period (Richter and Drayer, unpubl. data). In Maine, green frogs consumed all the wood frog egg masses (over 100) in two created pools in less than 1 week (Vasconcelos and Calhoun 2004). In short, failure of created pools to support viable populations of specialized pool breeders is often directly attributable to improper hydrological regime (Porej et al. 2004; Vasconcelos and Calhoun 2006; Gamble and Mitsch 2009; Drayer 2012; Denton and Richter 2013; Kross 2014).
Hydro-geomorphic setting
Vernal pools may develop in a variety of hydro-geomorphic settings including riparian areas, surface water depressions (often in the upper reaches of a watershed), and as groundwater depressions in low lying areas (either embedded in larger wetlands or as discrete depressions). Each of these settings is characterized by different hydro-periods (ranging from ephemeral to semi-permanent vernal pools) and major water inputs (e.g., precipitation, groundwater inflow, subsurface flow, and runoff) that may influence water temperature, pool substrate, and water chemistry (Whigham and Jordan 2003; Sacerdote and King 2009; Gebo and Brooks 2012).
For example, pools that receive inputs from groundwater as well as from surface runoff tend to have mucky soils and water chemistry that is less variable than pools supported primarily by surface water. Water chemistry and 8 temperature will affect rates of embryo and larval development in pools (Newman 1998; Karraker et al. 2008; Battaglin et al. 2009). Local populations of amphibians may be particularly adapted to these subtle differences in pool setting (Rice and Emery 2003; Brady 2012). However, there is no published information on how to create pools in these different landscape settings to ensure replication of the abiotic environment.
Vegetation
Within pool vegetation structure. In our region, the composition and structure of pool vegetation varies greatly, and only a few plant species are true vernal pool obligates (e.g., Scirpus ancistrochaetus [northeastern bulrush] and Hottonia inflata [feather-foil]; Cutko and Rawinski 2008; Drayer 2012; Denton and Richter 2013; Pennsylvania Natural Heritage Program 2013). Although vegetation in pools may be important for providing shade, refuge, and potential egg attachment sites (Porej and Hetherington 2005; Vasconcelos and Calhoun 2006; Skidds et al. 2007; King 2012), many pools may be largely devoid of vascular plants. In these cases, fine and coarse woody material deposited in pools from adjacent trees and shrubs may serve as egg attachment sites and refugia from predation.
Potential for invasive species. The potential for colonization by weedy plants in created pools is high, and pools can become dominated by monocultures of cattail (Typha spp.), common reed (Phragmites spp. ), purple loosestrife ( Lythrum salicaria ), or duckweed (Lemna spp.) in disturbed sites, especially in pools located in more open settings. These plants can eventually fill or dry out a vernal pool, dramatically changing the hydro-period and altering pool chemistry and temperature (Vasconcelos and Calhoun 2004; Cutko and Rawinski 2008).
Pool canopy cover. Canopy cover adjacent to pools affects pool temperature and supporting detrital food webs through inputs of leaf litter, which impacts developmental rates and species interactions (reviewed in deMaynadier and Hunter 1995; Schiesari 2006; deMaynadier and Houlahan 2008). Canopy removal directly over 9 and around vernal pools may alter amphibian community composition as increased light, temperature, and primary productivity may attract a broader array of amphibian and invertebrate species that either compete or prey on pool breeding specialists (Brooks et al. 1998; Calhoun and deMaynadier 2002; Skelly et al. 2002; Denton and Richter 2013). Thus, reduction in canopy cover may preferentially support species typically found in opencanopy wetlands to the detriment of vernal pool specialists. For example, Schiesari (2006) found that northern leopard frogs (Lithobates pipiens) outperformed wood frogs in open canopy, but not closed canopy conditions.
Additionally, pools may become dominated by algae and duckweed, ultimately reducing oxygen levels (e.g., by limiting light infiltration and reducing algal photosynthesis) and limiting their suitability for target pool breeders (Vasconcelos, unpublished data).
Slope
Most natural vernal pools have shallow littoral zones available with gradual slopes to the center, which has been linked to greater species richness compared to constructed wetlands (Porej and Hetherington 2005; Drayer 2012). In the ridgetop wetland ecosystem of Kentucky, Drayer (2012) found natural wetlands to have significantly lower slope (measured as depth at 1 m from shore; mean±1 SE=9.1±0.8 cm) than constructed wetlands (15.4±1.6 cm). Created pools that do not mimic the slope of natural pools may be less suitable for amphibian recruitment (Porej and Hetherington 2005; Croshaw and Scott 2006). Steep, abrupt slopes may cause access problems for salamanders and may limit the growth of vegetation (Simon et al. 2009; Shulse et al. 2012) or the rate of melting ice cover in northern climates. Additionally, shallow areas can be important for predator avoidance (Porej and Hetherington 2005), thermoregulation of amphibians for growth (Wellborn et al. 1996) and to decrease the occurrence of diseases (Raffel et al. 2010). Optimal slopes vary from pool to pool depending on typical levels of winter and fall rains (Croshaw and Scott 2006). Marbled salamanders are particularly sensitive to changes in pool slope because their selection of nesting sites depends on the interplay of pool bathymetry and water levels determined by winter and fall rains (Croshaw and Scott 2006). 10 Soils
Soil development is highly dependent upon hydro-period (i.e., longer hydro-periods result in saturated conditions that can create mucky, organic surface layers) and parent materials (e.g., highly permeable sandy outwash materials versus compact, restrictive tills; Tiner and Veneman 1989). If creation efforts properly mimic the necessary hydrology and pools are placed in appropriate hydro-geomorphic settings, soil development in created sites should be similar to that of natural pools (see Biebighauser 2003). If soils from the original wetland are excavated and used in the created wetland, this should enhance the likelihood of colonization by native species (i.e., by plant seeds and resting stages of invertebrates or dormant cysts or eggs; Stauffer and Brooks 1997; Colburn et al. 2008).
Soil compaction in created pools may influence hydro-period (Whittecar and Daniels 1999; Gamble and Mitsch 2009). If the soils have sufficient clay, then they can be readily compacted to hold water (Biebighauser 2003, 2011). However, there is always a risk that they will hold water too long (Korfel et al. 2009). For example in Kentucky, constructed wetlands with the same hydro-geomorphic parameters as natural wetlands and maximum depths less than 30 cm were permanently inundated, presumably because of soil compaction (Drayer 2012; Denton and Richter 2013).
Landscape setting and pool context affect success of creation projects Because of the complex life histories of pool breeding specialists, conservation efforts must support critical population dynamics including migration from winter hibernacula to breeding pools, migration from breeding pools to summer habitat, migration from summer refugia to hibernacula, and juvenile dispersal (leaving the natal pool to breed in new pools). Not all pool breeding amphibians have the same post breeding habitat 11 needs (see Baldwin et al. 2006a; Semlitsch et al. 2008; Rittenhouse and Semlitsch 2007). For example, wood frogs in the northeastern U.S. often require three distinct habitat elements including the breeding pool, summer refugia of forested wetlands or hillside seeps, and well drained upland forests for hibernation (Baldwin et al. 2006b, Groff, unpublished data). In contrast, salamanders may require fewer habitat elements, but may be more sensitive to habitat disturbance closer to the pool and to disruption of migration routes (McDonough Haughley and Paton 2007).
The literature emphasizes the importance of the forest matrix linking pools as post breeding habitat and as migratory and dispersal corridors (see Compton et al. 2007; Harper et al. 2008; Greenwald et al. 2009; Baldwin and deMaynadier 2009; Denton and Richter 2013; Peterman et al. 2013). Abundant coarse woody material and leaf litter on the forest floor and at least 50 % canopy cover have been suggested as key components of suitable forest habitat (Gibbs 1998; Patrick et al. 2006; deMaynadier and Houlahan 2008).
Disturbance thresholds for each species of pool breeding amphibian have not been widely investigated, but a number of case studies (Gibbs 1998; Homan et al. 2004; Windmiller et al. 2008; Windmiller and Calhoun 2008) and extinction models (Harper et al. 2008) suggest that current regulatory zones around pools in states with specific vernal pool legislation are inadequate for conserving pool breeding amphibian populations. Altering as little as 25 % of post breeding habitat adjacent to pools led to local extinctions in two Massachusetts vernal pools (Windmiller et al. 2008).
In addition to affecting upland habitat quality for amphibians, upland vegetation characteristics may influence pool characteristics at large spatial scales. Skelly et al. (1999) suggested that forest recovery following agricultural abandonment increased evapotranspiration rates, leading to decreased hydro-periods. Presumably, any changes to vegetation within the catchment area of a pool, such as those caused by succession or timber harvest, have the potential to affect a pool’s water balance (Skelly et al. 2005). 12 A rich body of literature has documented pool breeding amphibian migration and dispersal distances from breeding pools (Richter et al. 2001; Semlitsch 2002; Faccio 2003; Regosin et al. 2003; Cushman 2006; Baldwin et al. 2006b; Patrick et al. 2006; Rittenhouse and Semlitsch 2007, see review in Semlitsch and Skelly 2008). It is generally accepted that the area within 300 m of a breeding pool is core foraging and migration habitat for most adult pool breeding species (Rittenhouse and Semlitsch 2007; Harper et al. 2008), but this area can exceed 1,000 m (e.g., Humphries and Sisson 2012). Genetically significant linkages among pools typically occur in neighborhoods of pool based populations within 10 km of one another (see Smith and Green 2005; Gibbs and Reed 2008). Data on juvenile movement is limited because it is often not possible to conduct radio telemetry studies on animals this size (but see Roznik and Johnson 2009). Alternative methods (e.g., fluorescent powder tracking or pit tagging) only allow for short term assessment of microhabitat needs and movement pattern (Emanetoglu et al. 2013). Dispersing juveniles may travel in excess of 1 km from natal pools (Patrick et al. 2006; Gamble et al. 2007), if not significantly farther (Smith and Green 2005). Frogs typically travel farther than salamanders, and females generally travel farther than males (Regosin et al. 2003; Rittenhouse and Semlitsch 2007). Although it is not usually feasible for a mitigation project to control habitat quality at these distances, the constraints imposed by this larger landscape perspective should be recognized because landscape context is crucial to the functioning of pools (see Kentula 2000; Zedler 2000; Regosin et al. 2005; Julian 2009; Gardner et al. 2007) and to preserve local genetic diversity (Gibbs and Reed 2008; Richter et al. 2013b). Many researchers have documented that amphibians are particularly sensitive to pool context and quality of the surrounding habitat (Kolozsvary and Swihart 1999; Lehtinen et al. 1999; Guerry and Hunter 2002; Babbitt et al. 2009; Simon et al. 2009).
We know of only one peer reviewed journal article that provides prescriptive measures for the spatial configuration of pools at the landscape scale. Petranka and Holbrook (2006) urged restoration ecologists who are designing creation or restoration projects to consider whether natural pools are patchy (clustered, allowing free flow from pool to pool within a single amphibian population) or distant (suggesting a meta-population 13 structure with limited interaction among pools). If pools are patchy, they suggest that creation should provide clusters of pools with varying hydro-periods to support local anti-predator responses (i.e., allowing amphibians to switch to pools with less predation pressure, which can fluctuate annually). However, we note that “landscape” is a term that can be used at many scales and one needs to think about particular species to assess whether pools are “patchy” or “distant.” For example, are pools within genetic neighborhoods of dispersing amphibians, and are there significant barriers interrupting typical movement patterns? Emulating natural patterns of pool distribution and post-breeding habitat may be difficult depending on the extent and spatial pattern of the disturbance that initiated the mitigation effort, availability of suitable terrestrial habitat and suitable pool sites, and access to information on historical patterns of pools or other wetlands. For example, Denton and Richter (2013) and Drayer (2012) compared created and natural wetlands on ridgetops in Kentucky’s Daniel Boone National Forest and found that the creation of permanent pools over the last 20 years has likely provided avenues of dispersal and migration for green frogs, bullfrogs, and eastern newts, thereby potentially exposing naturally occurring ridge top amphibian species to direct predation and to diseases such as Batrachochytrium dendrobatidis (amphibian chytrid fungus) and ranavirus (Daszak et al. 2004; Gahl et al. 2009; Greenspan et al. 2012; Richter et al. 2013a). In that landscape, permanent created pools did not support wood frogs and marbled salamanders; those species remained restricted to natural, ephemeral pools.
Summary
Collectively, the peer reviewed literature cautions against assuming pool creation will successfully create suitable habitat for target vernal pool species. It suggests that the key to effective vernal pool creation is attention to context: what was the historical landscape distribution of wetlands and vernal pools and what is the current distribution? Seen through the lens of past and present wetland status, the sometimes blurred or conflicting messages from the current technical and scientific literature come into focus. For example, in some 14 cases, where natural pools have been highly degraded or locally extirpated, creation may be the only option in efforts to support wetland dependent wildlife. This perspective underlies the work of Biebighauser (2003) whose guidance for creating pools was developed in a region with limited wetland resources where providing habitat for any wetland wildlife is likely to be judged desirable (also see Brand and Snodgrass 2009; Simon et al. 2009; Brown et al. 2012). However, such creation efforts carry a high risk of failure for vernal pool species that are sensitive to predators associated with more permanent waters and those that need adjacent forest for postbreeding habitat (e.g., wood frogs) (Boone et al. 2008; Drayer 2012; Denton and Richter 2013; Kross 2014).
Our review of the literature indicates that vernal pool creation is an imperfect science and should be used as a last resort after exhausting more reliable protective methods (Calhoun et al. 2005; Windmiller and Calhoun 2008; Denton and Richter 2013). The practice is perhaps appropriate in landscapes that have been subjected to severe wetland losses, such as former agricultural landscapes where forests have recovered but drained and destroyed wetlands have not. Vernal pool ecosystems are difficult candidates for creation because the community structure is as tied to the surrounding forested ecosystems as to the actual pool depression and because pool function is so tightly tied to hydrology. In addition, ideal breeding site characteristics may vary among pool breeding species regionally (Snodgrass et al. 2000; Petranka et al. 2007) as do post breeding habitat quality tolerances (Windmiller et al. 2008). For these reasons, mitigation efforts must, first and foremost, consider conserving existing pools in a suitable landscape, and, if that is impossible, seek to emulate pools in the region in terms of hydro-geomorphic setting, spatial distribution, and natural amphibian communities.
Recommendations
Pool specific
If pool hydro-period is successfully recreated, generally the rest will follow if source populations exist within 15 viable colonization distances. However, hydro-period is perhaps the most difficult parameter to reproduce.
Therefore, we caution practitioners to carefully consider the effects of substrate composition, compaction, maintenance of canopy, and construction on pool fauna when building vernal pools. If pool hydrology is permanent or even semi-permanent with connections to other waters, fish and amphibian species typically restricted to permanent wetlands may colonize the pool and decimate vernal poolbreeding amphibians by depredating both egg masses and larvae. If the vegetation structure is complex, some ambystomatid species may be able to breed successfully for a while, but evidence suggests that wood frogs will not persist (Vasconcelos and Calhoun 2006; Denton and Richter 2013; Julian et al. 2013; Kross 2014). We recommend that practitioners create pools that support appropriate, natural vegetation (Pearl et al. 2005), hydro-geomorphic setting, and soil type and compaction as much as possible. Limiting removal of adjacent trees and root damage during construction will preserve canopy cover, and reducing soil compaction will facilitate colonization by herbaceous plants and trees. Obviously, reestablishment of mature forest, which is often the dominant cover associated with vernal pools in this region, may take many decades. Wood frogs and ambystomatid species will use shrubs and persistent graminoids for egg attachment sites.
Because littoral zones with a gradual slope are important for life history requirements of some species (e.g., marbled salamanders) and provide more opportunity for establishment of vegetation, thermoregulation of larvae, and refuge from predation, we recommend emulating slopes of local, natural pools.
Landscape and population scale considerations
We recommend paying attention to local landscape context and desired endpoints of mitigation. Pool breeding amphibians may be organized in meta-populations that function through exchanges among pools following catastrophic events (rescue and colonization processes) and that depend upon relatively long distance dispersal events to maintain genetic integrity (Gibbs and Reed 2008). Alternatively, pool breeding 16 amphibians may occur in patchy populations when pools are relatively close (<500 m) making pools of varying hydro-period in close proximity important for allowing natural exchanges of breeding adults in responses to within pool habitat conditions (e.g., responding to changing predator densities in pools; see Petranka and Holbrook 2006).
We advise against creating permanent pools. Addition of created pools with permanent hydro-periods, especially at unnaturally high densities, may subsidize local populations of bullfrogs, green frogs, and eastern newts. This can negatively affect nearby pool breeding specialists by increasing predation pressure on eggs and larvae and by increasing the likelihood of disease introductions (Gahl et al. 2009; Brown et al. 2012; Greenspan et al. 2012; Denton and Richter 2013), thus potentially creating population sinks. This possibility argues against the strategy of creating an excess number of vernal pools in the hope that some portion of them will have the correct hydro-period.
We recommend ample post breeding habitat and pool connections be maintained for long term amphibian population persistence. The 30 m surrounding pools often serves as critical habitat for newly emerged amphibians and should provide ample canopy cover, leaf litter, and coarse woody material for cover and foraging (Regosin et al. 2003; deMaynadier and Houlahan 2008). However, persistence of populations is favored by at least 300 m of forested habitat and connections that allow movements among pools and pool clusters.
Measures of success over the monitoring period
Presence of an amphibian species is a poor indicator of population health; individuals may be transient animals that are no longer able to breed successfully in the pool. Trends in abundance of target species over at least 5 years post creation is a better measure of population robustness, although even abundance may not reflect breeding success in the pool. Ideally, the best measure of success is evidence of breeding and metamorphosis by intended species over five or more years. 17 We recommend a minimum of five years of post-creation monitoring reproductive effort of target species (number of egg masses laid) and recruitment (number of animals successfully metamorphosing and returning to breed). Seemingly functional pools may be ecological traps where individuals breed but where larvae are not able to mature and leave the pool; thus recruitment is a superior metric to egg mass counts.
Survivorship to metamorphosis is a better predictor of population health than status of eggs or larvae (Richter et al. 2003; Cushman 2006; Rittenhouse and Semlitsch 2007; Harper et al. 2008). Similarly, because it may take a number of years for invasive vegetation to occupy and dominate a pool, we recommend monitoring pool vegetation for at least five years after creation. We do not recommend using species richness as an indicator of pool success (or the contribution of that pool to regional biodiversity). In fact, richness can be similar between constructed and natural pools across the same landscape, while community composition is different, driven by the presence of species not historically breeding in the ecosystem (Drayer 2012; Denton and Richter 2013). Wood frogs and most of the region’s ambystomatid salamanders are forest pool specialists, in part because shaded pools are less likely to harbor competitors and predators including green frogs, bullfrogs, northern leopard frogs, pickerel frogs (Lithobates palustris), and eastern newts (Korfel et al. 2009).
We strongly discourage using occupation by widespread, generalist species such as green frogs and their tadpoles as a criterion for success as such species do not adequately replace pool functions or contribute to landscape scale biodiversity. In contrast, these species are an indicator of failure to replace pool functions, especially for wood frogs.
We also recommend monitoring from a long term perspective because natural vernal pools tend towards extreme variability, from pool to pool and even within the same pool from year to year. Just as a natural vernal pool may not be productive or “successful” in a given year, a similar allowance should be considered in evaluating created pools. Factors such as weather, ecological context, and the developmental and demographic needs of target species can inform locally appropriate success criteria. Indeed, we hypothesize that having occasional years where pool hydro-period is very short may facilitate amphibian reproductive 18 success over the long run by both limiting disease and reducing populations of predatory invertebrates.
Adaptive management
We recommend an adaptive approach (see Salafsky et al. 2001) to management of vernal pools, in which scientists and practitioners collaborate. First and foremost, it should be determined if there is a need for vernal pool creation based on the historical distribution and abundance of vernal pools and other wetlands, the loss and degradation of these resources, and the desired future condition. This discussion should include diverse stakeholders (e.g., landowners, government officials, scientists, practitioners, and others) so that the best management policy for a particular region can be determined (Jansujwicz et al. 2013a, 2013b). Following this, management protocols (i.e., construction techniques) should be developed by key technical stakeholders including scientists, engineers, and managers. Scientists should then monitor the success of pool creation based on multiple indicators, including hydro-period and successful reproduction by focal amphibian species. These results should be used to refine approaches to construction, if necessary. As part of the iterative process, research might be necessary to determine the appropriate construction techniques based on native soils, hydrogeographic characteristics, and other factors affecting hydrology (Brown and Richter 2012; Drayer 2012; Denton and Richter 2013).
Future Research Needs
Even if practitioners consult the papers discussed in this review before embarking on mitigation activities, they offer little technical guidance, especially in terms of assuring ecologically successful outcomes. Below are research needs that became obvious as a result of our review:
Regional studies are needed to determine appropriate compaction parameters based on ambient soil texture, water source (surface/precipitation versus groundwater), and geomorphology of proposed 19 construction sites. The interaction of these parameters should be considered relative to resulting hydroperiods as well as biological outcomes. Post-breeding habitat requirements of vernal pool species need to be better defined across their ranges to inform landscape quality and quantity issues such as directing vernal pool creation projects to landscape contexts in which they have the greatest probability of success. Soil and substrate structure of vernal pools needs further study so that the implications of using compacted soil to retain water is balanced against other life history needs. Longer term studies will reveal the permanence of outcomes observed within typical monitoring periods (<5 years) and if successional trajectories in created pools differ from natural pools, which are themselves poorly studied in this regard. Rapidly changing climatic conditions may also necessitate reevaluation of earlier research, as well as present opportunities to design pools to accommodate predicted effects of climate change on wetland-dependent wildlife.
Reporting outcomes, both positive and negative, is the only means by which the practice of vernal pool creation will progress. Whether in project reports or peer reviewed articles, including sufficient detail about pool design parameters and landscape context is essential to guide future projects as well as research, especially when these characteristics can be correlated with outcomes. 20 Acknowledgments
This is Maine Agriculture and Forestry Experiment Station Publication 3371 21 Figures
Figure 1
Amphibian species use vernal pools for oviposition and larval rearing according to different schedules throughout the year. Typical species such as wood frogs and spotted salamanders migrate to pools in late winter or early spring, but larval development takes several weeks longer for spotted salamanders than for wood frogs. In contrast, marbled salamanders migrate to pools in late summer and deposit eggs in terrestrial nest sites within the dry basin. Embryos hatch following autumn inundation, and larvae overwinter and metamorphose on a schedule similar to wood frogs.
Data are based on movement phenology of adults and metamorphosed amphibians reported by Paton and Crouch (2002) in Rhode Island, U.S.A.; shading gradient between endpoints and black interior reflects 5-95 % range of cumulative percentiles of number of migrants. Changes in latitude, altitude, and weather will influence end points. For example, Guttman et al. (1991) estimated that wood frog breeding commences a mean of 5.2 days later per additional degree of latitude
22 References
Babbitt KJ, Baber MJ, Tarr TL (2003) Patterns of larval amphibian distribution along the wetland hydro-period gradient. Can J Zool 81:1539–1552
Babbitt KJ, Baber MJ, Childers DL, Hocking D (2009) Influence of agricultural upland habitat type on larval anuran assemblages in seasonally inundated wetlands. Wetlands 29:294–301 Baldwin RF, deMaynadier PG (2009) Assessing threats to pool breeding amphibian habitat in an urbanizing landscape. Biol Conserv 142: 1628–1638
Baldwin RF, Calhoun AJK, deMaynadier PG (2006a) The significance of hydro-period and stand maturity for pool breeding amphibians in forested landscapes. Can J Zool 84:1604–1615 Baldwin RF, Calhoun AJK, deMaynadier PG (2006b) Conservation planning for amphibian species with complex habitat requirements: a case study using movements and habitat selection of the wood frog Rana sylvatica. J Herpetol 40:443–454
Battaglin WA, Rice KC et al (2009) The occurrence of glyphosate, atrazine, and other pesticides in vernal pools and adjacent streams in Washington, DC, Maryland, Iowa, and Wyoming, 2005–2006. Environ Monit Assess 155:281–307 Biebighauser TR (2003) A guide to creating vernal ponds. USDA Forest Service, South Morehead Biebighauser TR (2011) Wetland restoration and construction: a technical guide. Upper Susquehanna Coalition, Owego Black C, Zedler PH (1996) An overview of 15 years of vernal pool restoration and construction activities in San Diego County, California. In: Witham CW, Bauder ET, Belk D, Ferren WR Jr, Ornduff R (eds) Ecology, conservation, and management of vernal pool ecosystems – proceedings from a 1996 conference. California Native Plant Society, Sacramento, pp 195–205
Boone MD, Semlitsch RD, Mosby C (2008) Suitability of golf course ponds for amphibian metamorphosis when bullfrogs are removed. Conserv Biol 22:172–179
Brady SP (2012) Road to evolution? Local adaptation to road adjacency in an amphibian (Ambystoma maculatum). Scientific Reports 2 Brainwood M, Burgin S (2009) Hotspots of biodiversity or homogeneous landscapes? Farm dams as biodiversity reserves in Australia. Biodivers Conserv 18:3043–3052 Brand AB, Snodgrass JW (2009) Value of artificial habitats for amphibian reproduction in altered landscapes. Conserv Biol 24:295–301
Brooks RP, Gebo NA (2013) Wetlands mitigation and restoration. In: Brooks RP, Wardrop DH (eds) Mid Atlantic freshwater wetlands: advances in wetlands science, management, policy, and practice. Springer Science+Business Media, NY, pp 421–440
Brooks RT, Stone J, Lyons P (1998) An inventory of seasonal forest ponds on the Quabbin reservoir watershed, Massachusetts. Northeast Nat 5:219–230
Brown DR, Richter SC (2012) Meeting the challenges to preserving Kentucky’s biodiversity. Sustainability 25:22–33 Brown DJ, Street GM, Nairn RW, Forstner MRJ (2012) A place to call home: amphibian use of created and restored wetlands.
Article 989872: International Journal of Ecology
Calhoun AJK, deMaynadier P (2002) Forestry habitat management guidelines for vernal pool wildlife in Maine. Wildlife 23 Conservation Society Technical Paper #6 Rye, NY
Calhoun AJK, deMaynadier P (2008) Science and conservation of vernal pools in northeastern North America. CRC Press, Boca Raton, Florida Colburn EA (2004) Vernal pools: natural history and conservation. McDonald and Woodward Publishing Company, Blacksburg, VA
Calhoun AJK, Miller NA, Klemens MW (2005) Conserving poolbreeding amphibians in human dominated landscapes through local implementation of best development practices. Wetl Ecol Manag 13: 291–304 Colburn E, Weeks SC, Reed SK (2008) Diversity and ecology of vernal pool invertebrates. In: Calhoun AJK, deMaynadier PG (eds) Science and conservation of vernal pools in Sortheastern North America. CRC Press, Boca Raton, pp 107–124 Cole CA, Brooks RP (2000) Patterns of wetland hydrology in the ridge and valley province, Pennsylvania, USA. Wetlands 20:
438–447
Compton BW, McGarigal K, Cushman SA, Gamble LR (2007) A resistant kernel model of connectivity for amphibians that breed in vernal pools. Conserv Biol 21:788–799
Council NR (1992) Restoration of aquatic ecosystems. National Academy Press, Washington Council NR (2001) Compensating for wetland losses under the clean water Act. National Academy Press, Washington Croshaw DA, Scott DE (2006) Marbled salamanders (Ambystoma opacum) choose low elevation nest sites when cover availability is controlled. Amphibia Reptilia 27:359–364 Cushman SA (2006) Effects of habitat loss and fragmentation on amphibians: a review and prospectus. Biol Conserv 128:
231–240
Cutko A, Rawinski RJ (2008) Flora of Northeastern vernal pools. In: Calhoun AJK, deMaynadier PE (eds) Science and conservation of vernal pools in Northeastern North America. CRC Press, Boca Raton, pp 71–104 da Silva FR, Gibbs JP, Rossa Feres DD (2011) Breeding habitat and landscape correlates of frog diversity and abundance in a tropical agricultural landscape. Wetlands 31:1079–1087
Dahl TE (2011) Status and trends of wetland in the conterminous United State 2004 to 2009. U.S. Department of the Interior, Fish and Wildlife Service, Washington
Daszak P, Cunningham AA, Loncore JE, Brown C, Porter D (2004) Experimental evidence that the bullfrog (Rana catesbeiana) is a potential carrier of chytridiomycosis, and emerging fungal disease of amphibians. Herpetol J 14:201–207 deMaynadier PG, Hunter ML Jr (1995) The relationship between forest management and amphibian ecology: a review of the North American literature. Environ Rev 3:230–261
deMaynadier PG, Hunter ML Jr (1999) Forest canopy closure and juvenile emigration by pool breeding amphibians in Maine.
J Wildl Manag 63:441–450
Denton RD, Richter SC (2013) Amphibian communities in natural and constructed ridge top wetlands with implications for wetland construction. J Wildl Manag 77:886–889
Dimauro D, Hunter ML Jr (2002) Reproduction of amphibians in natural and anthropogenic temporary pools in managed forests. For Sci 48: 397–406
Drayer AN (2012) Efficacy of constructed wetlands of various depths for natural amphibian community conservation. Thesis, Eastern Kentucky University, KY
Emanetoglu NW, Aumann HM, Kus EM, Cline BB (2013) A low cost harmonic radar for tracking very small tagged 24 amphibians. IEEE (Institute of Electrical and Electronics Engineers) Transactions on Instrumentation and Measurement
Faccio SD (2003) Postbreeding emigration and habitat use by Jefferson and spotted salamanders in Vermont. J Herpetol 37:479–489
Gahl MK, Calhoun AJK, Graves R (2009) Facultative use of seasonal pools by American bullfrogs (Rana catesbeiana).
Wetlands 29: 697–703
Gahl MK, Longcore JE, Houlahan JE (2012) Varying responses of northeastern North American amphibians to the chytrid pathogen Batrachochytrium dendrobatidis. Conserv Biol 26: 135–141Gamble DL, Mitsch WJ (2009) Hydro-periods of created and natural vernal pools in central Ohio: a comparison of depth and duration of inundation. Wetl Ecol Manag 17:385–395
Gamble LR, McGarigal K, Compton BW (2007) Fidelity and dispersal in the pond breeding amphibian, Ambystoma opacum:
implications for spatio temporal population dynamics and conservation. Biol Conserv 139:247–257 Gardner TA, Barlow J, Peres CA (2007) Paradox, presumption, and pitfalls in conservation biology: the importance of habitat change for amphibians and reptiles. Biol Conserv 138:166–179 Gebo NA, Brooks RP (2012) Hydro-geomorphic (HGM) assessments of mitigation sites compared to natural reference wetlands in Pennsylvania. Wetlands 32:321–331
Gibbs JP (1998) Distribution of woodland amphibians along a forest fragmentation gradient. Landsc Ecol 13:263–268 Gibbs JP, Reed JM (2008) Population and genetic linkages of vernal pool associated amphibians. In: Calhoun AJK, deMaynadier PG (eds) Science and conservation of vernal pools in northeastern North America. CRC Press, Boca Raton, pp 149–167
Good CD (2006) Constructed ponds as mitigated habitat for the wood frog (Rana sylvatica LeConte) and the spotted salamander (Ambystoma maculatum Shaw) in West Virginia. Thesis, Marshall University, WV Gray MJ, Miller DL, Hoverman JT (2009) Ecology and pathology of amphibian ranaviruses. Dis Aquat Org 87:243–266 Greenspan SE, Calhoun AJK, Loncore JE, Levy MG (2012) Transmissions of Batrachochytrium dendrobatidis to wood frogs (Lithobates sylvaticus) via a Bullfrog (L. catesbeianus) vector. J Wildl Dis 48:575–582 Greenwald KR, Gibbs HL, Waite TA (2009) Efficacy of land cover models in predicting isolation of marbled salamander populations in a fragmented landscape. Conserv Biol 23:1232–1241 Guerry AD, Hunter ML Jr (2002) Amphibian distributions in a landscape of forests and agriculture: an examination of landscape composition. Conserv Biol 16:745–754
Guttman SI, Bramble JE, Sexton OJ (1991) Observations on the breeding immigration of wood frogs Rana sylvatica reintroduced in eastcentral Missouri. Am Midl Nat 125:269–274 Harper EB, Rittenhouse TAG, Semlitsch RD (2008) Demographic consequences of terrestrial habitat loss for pool breeding amphibians: predicting extinction risks associated with inadequate size of buffer zones. Conserv Biol 22:1205–1215 Hoeltje SM, Cole CA (2007) Losing function through wetland mitigation in central Pennsylvania, USA. Environ Manag 39:385–402
Homan RN, Windmiller BS, Reed JM (2004) Critical thresholds associated with habitat loss for two vernal pool breeding amphibians. Ecol Appl 14:1547–1553
25 Humphries WJ, Sisson MA (2012) Long distance migrations, landscape use, and vulnerability to prescribed fire of the gopher frog (Lithobates capito). J Herpetol 46:665–670
Jansujwicz JS, Calhoun AJK, Leahy JE, Lilieholm RJ (2013a) Using mixed methods to develop a frame based private landowner typology. Soc Nat Resour Int J. doi:10.1080/08941920.2012.729294 Jansujwicz JS, Calhoun AJK, Lilieholm R (2013b) The Maine vernal pool mapping and assessment program: engaging municipal officials and private landowners in community based citizen science. Environ Manag. doi:10.1007/s00267 013 0168
Julian JT (2009) Evaluating amphibian occurrence models and the importance of small, isolated wetlands in the Delaware Water Gap National Recreational Area. Dissertation, Pennsylvania State University, PA Julian JT, Rocco G, Turner MM, Brooks RP (2013) Assessing wetlandriparian amphibian and reptile communities of the Mid Atlantic region. In: Brooks RP, Wardrop DH (eds) Mid Atlantic freshwater wetlands: advances in wetlands science, management, policy, and practice. Springer Science and Business Media, NY, pp 313–337 Karraker NE, Gibbs JP, Vonesh JR (2008) Impacts of road deicing salt on the demography of vernal pool breeding amphibians. Ecol Appl 18: 724–734
Kentula ME (2000) Perspectives on setting success criteria for wetland restoration. Ecol Eng 15:199–209 Kentula ME, Brooks RP, Gwin SE, Holland CC, Sherman AD, Sifneos JC (1992) An approach to improving decisionmaking in wetland restoration and creation. Island Press, Washington King SK (2012) Four toed salamander (Hemidactylium scutatum) nest site characteristics in natural and constructed wetlands in eastern Kentucky. Thesis, Eastern Kentucky University Kolozsvary MB, Swihart RK (1999) Habitat fragmentation and the distribution of amphibians: patch and landscape correlates in farmland. Can J Zool 77:1288–1299
Korfel CA, Mitsch WJ, Hetherington TE, Mack JJ (2009) Hydrology, physiochemistry, and amphibians in natural and created vernal pool wetlands. Restor Ecol 18:843–854
Kross, CS (2014) Interactions between eastern newts and wood frogs in an anthropogenically altered wetland system in eastern Kentucky. Thesis, Eastern Kentucky University, KY Kusler JS, Kentula ME (1990) Wetland creation and restoration: the status of the science. Island, Washington Lehtinen RM, Galatowitsch SM, Tester JR (1999) Consequences of habitat loss and fragmentation for wetland amphibian assemblages. Wetlands 19:1–12
Leibowitz SG (2003) Isolated wetlands and their functions: an ecological perspective. Wetlands 23:517–531 Lichko LE, Calhoun AJK (2003) An evaluation of vernal pool creation projects in New England: project documentation from 1991–2000. Environ Manag 32:141–151
Magnusson AK, Williams DD (2006) The roles of natural temporal and spatial variation versus biotic influences in shaping the physicochemical environment of intermittent ponds: a case study. Arch Hydrobiologia 165:537–556 Mahaney WS, Klemens MW (2008) Vernal pool conservation policy: the federal, state, and local context. In: Calhoun AJK, deMaynadier PG (eds) Science and conservation of vernal pools in northeastern North America. CRS Press, Boca Raton, pp 19–212
McDonough Haughley C, Paton PWC (2007) Spotted salamander dispersal across a forested landscape fragmented by a golf 26 course. J Wildl Manag 71:1163–1169
Moreno Mateos D, Power ME, Comin FA, Yockteng R (2012) Structural and functional loss in restored wetland ecosystems.
PLoS Biol 10: e1001247
Newman RA (1998) Ecological constraints on amphibian metamorphosis: interactions of temperature and larval density with responses to changing food level. Oecologia 115:9–16
Palik B, Batzer DP, Kern C (2006) Upland forest linkages to seasonal wetlands: litter flux, processing, and food quality.
Ecosystems 9: 142–151
Paton PWC, Crouch WB III (2002) Using the phenology of pondbreeding amphibians to develop conservation strategies.
Conserv Biol 16:194–204
Patrick DA, Hunter ML Jr, Calhoun AJK (2006) Effects of experimental forestry treatments on a Maine amphibian community.
For Ecol Manag 234:323–332
Pearl CA, Adams MJ, Leuthold N, Bury RB (2005) Amphibian occurrence and aquatic invaders in a changing landscape:
implications for wetland mitigation in Willamette Valley, Oregon, USA. Wetlands 25:76–88 Pennsylvania Natural Heritage Program (2013) http://www. naturalheritage.state.pa.us/factsheets/15236.pdf accessed 6 June 2013
Peterman WE, Feist SM, Semlitsch RD, Eggert LS (2013) Conservation and management of peripheral populations: spatial and temporal influences on the genetic structure of wood frog (Rana sylvatica) populations. Biol Conserv 158:351– 358
Petranka JW, Holbrook CT (2006) Wetland restoration for amphibians: should local sites be designed to support metapopulations or patchy populations? Restor Ecol 14:404–411 Petranka JW, Murray SW, Kennedy CA (2003) Response of amphibians to restoration of a southern Appalachian wetland:
perturbations confound post restoration assessment. Wetlands 23:278–290 Petranka JW, Harp EM, Holbrook CT, Hamel JA (2007) Long term persistence of amphibian populations in a restored wetland complex. Biol Conserv 138:371–380
Porej D, Hetherington TE (2005) Designing wetlands for amphibians: the importance of predatory fish and shallow littoral zones in structuring amphibian communities. Wetl Ecol Manag 13:445–455 Porej DP, Micacchion M, Hetherington TE (2004) Core terrestrial habitat for conservation of local populations of salamanders and wood frogs in agricultural landscapes. Biol Conserv 120:399–409 Raffel TR, Michel PJ, Sites EW, Rohr JR (2010) What drives chytrid infections in newt populations? Associations with substrate, temperature, and shade. EcoHealth 7:526–536
Regosin JV, Windmiller BS, Reed JM (2003) Terrestrial habitat use and winter densities of the wood frog (Rana sylvatica). J Herpetol 37: 390–394
Regosin JV, Windmiller BS, Homan RN, Reed JM (2005) Variation in terrestrial habitat use by four pool breeding amphibian species. J Wildl Manag 69:1481–1493
Rice KJ, Emery NC (2003) Managing microevolution: restoration in the face of global change. Front Ecol Environ 1:469–478 Richter SC, Young JE, Seigel RA, Johnson GN (2001) Post breeding movements of the dark gopher frog, Rana sevosa goin and netting: implications for conservation and management. J Herpetol 35:316–321 27 Richter SC, Young JE, Johnson GN, Seigel RA (2003) Stochastic variation in reproductive success of a rare frog, Rana sevosa:
implications for conservation and for monitoring amphibian populations. Biol Conserv 111:171–177 Richter SC, Drayer AN, Strong JR, Kross CS, Miller DL, Gray MJ (2013a) High prevalence of ranavirus infection in permanent constructed wetlands in eastern Kentucky, USA. Herpetol Rev 44:464–466 Richter SC, Price SJ, Kross CS, Alexander JR, Dorcas ME (2013b) Upland habitat quality and historic landscape composition Influence genetic variation of a pond breeding salamander. Diversity 5:724–733 Rittenhouse TAG, Semlitsch RD (2007) Postbreeding habitat use of wood frogs in a Missouri Oak Hickory forest. J Herpetol 41: 645–653
Roznik EA, Johnson SA (2009) Canopy closure and emigration by juvenile gopher frogs. J Wildl Manag 73:260–268 Ruhi A, Sebastian OS, Feo C, Franch M, Gascon S, Richter Boix A, Bois D, Llorente G (2012) Man made Mediterranean temporary ponds as a tool for amphibian conservation. Annales de Limnologie Int J Limnol 48:81–93 Sacerdote AB, King RB (2009) Dissolved oxygen requirements for hatching success of two ambystomatid salamanders in restored ephemeral ponds. Wetlands 29:1202–1213
Salafsky N, Marogoluis R, Redford K (2001) Adaptive management: a tool for conservation practitioners. Washington, DC Schiesari L (2006) Pond canopy cover: a resource gradient for anuran larvae. Freshw Biol 51:412–423 Seigel RA, Dinsmore A, Richter SC (2006) Artificial respiration for ponds: experimental use of well water to increase hydroperiod as a management option for temporary pond breeding amphibians. Wildl Soc Bull 34:1022–1027 Semlitsch RD (1998) Biological delineation of terrestrial buffer zones for salamanders. Conserv Biol 12:1113–1119 Semlitsch RD (2002) Critical elements for biologically based recovery plans for aquatic breeding amphibians. Conserv Biol 16:
619–629
Semlitsch RD (2008) Moving wetland mitigation towards conservation banking. Nati Wetl Newsl 30:16 Semlitsch RD, Skelly DK (2008) Ecology and conservation of poolbreeding amphibians. In: Calhoun AJK, deMaynadier PG (eds) Science and conservation of vernal pools in Northeastern North America. CRC Press, Boca Raton, pp 127–148 Semlitsch RD, Todd BD, Blomquist SM et al (2008) Effects of timber harvest on amphibian populations: understanding mechanisms from forest experiments. Bioscience 59:853–862 Shulse CD, Semlitsch RD, Trauth KM, Gardner JE (2012) Testing wetland features to increase amphibian reproductive success and species richness for mitigation and restoration. Ecol Appl 22: 1675–1688 Simon JA, Snodgrass JW, Casey RE, Sparling DW (2009) Spatial correlates of amphibian use of constructed wetlands in an urban landscape. Landsc Ecol 24:361–373
Skelly DK, Werner EE, Cortwright SA (1999) Long term distributional dynamics of a Michigan amphibian assemblage. Ecology 80:2326–2337
Skelly DK, Freidenburg LK, Kiesecker JM (2002) Forest canopy and the performance of larval amphibians. Ecology 83:983– 992
Skelly DK, Halverson AM, Freidenburg LK, Urban MC (2005) Canopy closure and amphibian diversity in forested wetlands.
Wetl Ecol Manag 13:261–268
Skidds DE, Golet FC, Paton PWC, Mitchell JC (2007) Habitat correlates of reproductive effort in wood frogs and spotted salamanders in an urbanizing watershed. J Herpetol 41:439–450 28 Smith AM, Green DM (2005) Dispersal and the metapopulation paradigm in amphibian ecology and conservation: are all amphibian populations metapopulations? Ecography 28:110–128 Snodgrass JW, Domoroski MJ, Bryan AL Jr, Burger J (2000) Relationships among isolated wetland size, hydro-period, and amphibian species richness: implications for wetland regulations. Conserv Biol 14:414–419 Stauffer AL, Brooks RP (1997) Plant and soil responses to salvaged marsh surface and organic matter amendments at a created wetland in central Pennsylvania. Wetlands 17:90–105 Timm BC, McGarigal K, Gamble LR (2007) Emigration timing of juvenile pond breeding amphibians in western Massachusetts. J Herpetol 41:243–250
Tiner RW, Veneman PLM (1989) Hydric soils of New England. University of Massachusetts Cooperative Extension Communication Center, Amherst
U.S. Environmental Protection Agency (2008) Compensatory mitigation for losses of aquatic resources. Fed Regist 73:19594–19705
Vasconcelos D, Calhoun AJK (2004) Movement patterns of adult and juvenile wood frogs (Rana sylvatica) and spotted salamanders (Ambystoma maculatum) in three restored vernal pools. J Herpetol 38:551–561 Vasconcelos D, Calhoun AJK (2006) Monitoring created seasonal pools for functional success: a six year case study of amphibian responses, Sears Island, Maine, USA. Wetlands 26:992–1003 Wellborn GA, Skelly DK, Werner EE (1996) Mechanisms creating community structure across a freshwater habitat gradient.
Ann Rev Ecol Syst 27:337–363
Whigham DF, Jordan TE (2003) Isolated wetlands and water quality.
Wetlands 23:541–549
Whittecar R, Daniels WL (1999) Use of hydro-geomorphic concepts to design created wetlands in southeastern Virginia.
Geomorphology 31:355–371
Williams P, Whitfield M, Biggs J (2008) How can we make new ponds biodiverse? A case study monitored over 7 years.
Hydrobiologia 597:137–148
Windmiller B, Calhoun AJK (2008) Conserving vernal pool wildlife in urbanizing landscapes. In: Calhoun AJK, deMaynadier PG (eds) Science and conservation of vernal pools in northeastern North America. CRC Press, Boca Raton, pp 233–251 Windmiller BS, Homan RN, Regosin JV, Willitts LA, Wells DL, Reed JM (2008) Two case studies of declines in vernal pool breeding amphibian populations following loss of adjacent upland forest habitat. Urban Herpetology.
Herpetological Conservation, Volume 3, Society for the Study of Amphibians and Reptiles, Salt Lake City, Utah Zedler JB (2000) Progress in wetland restoration ecology. Trends Ecol Evol 15:402–407 See discussions, stats, and author profiles for this publication at: https://www.researchgate.net/publication/242242679 Reproductive performance, age structure, and natal dispersal of Swainson's hawks in the Butte Valley, California
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REPRODUCTIVE PERFORMANCE, AGE STRUCTURE,
AND NATAL DISPERSAL OF SWAINSON'S
HAWKS IN THE BUTTE VALLEY, CALIFORNIA
BRIAN WOODBRIDGE
USDA Forest Service, Klamath National Forest, 1312 Fairlane Road, Yreka, CA 96097 U.S.A.
KAREN K. FINLEY
31338 S.W. Bellfountain Road, Corvallis, OR 97333 U.S.A.
PETER H. BLOOM
Western Foundation of Vertebrate Zoology, 439 Calle San Pablo, Camarillo, CA 93070 U.S.A.
ABSTRACT.--We monitored annual occupancy, reproductive prformance, and natal dispersal ofa marked population ofSwainson's hawks (Buteo swainsoni) from 1984-94. Annual territory occupancy varied from 61-90%. Mean annual nest success was 65% (SE = 3.4%), and annual fledging rate was 1.53 (SE --
0.14) young per nest attempt. Of 567 Swainson's hawks banded as nestlings during this study, 41 were later recaptured asbreeding adults. Mean age at recapture was 5.9 yr (SE • 0.37; range = 3-15 yr. The mean age of color-marked aults observed in either 1993 or 1994 was 8.2 yr (SE -- 0.52), and ranged from 4-15 yr. Dispersal distances from natal site to subsequent breeding site ranged from 0-18.1 km (mean = 8.2 km, SD = 3.1), and was not different from random distances among territories.
KEY WORDS: Buteo swainsoni; dispersal; mark-recapture; population demography; Swainsoh's hawk; territory occupancy.
Caracterlsticas reproductivas, estructura deedad y dispersion natal Buteo swainsoni en el Butte Valley, California
RESUMEN.--Desde 1984 a 1994, monitoreamos la ocupaci6n aual, caracterlsticas reproductivas y dispersi6n natal de una poblaci6n marcada de Buteo swainsoni. La ocupaci6n de territorio vari6 anualmente desde un 61 a un 90%. La media anual de 6xito del nido fue de un 65% (EE = 3.4). La tasa anual de polluelos fue de 1.53 (EE = 0.14)juveniles por nidificaci6n. De 567 individuos marcados como polluelos duranteste estudio, 41 fueron recapturados como adultos reproductivos. La edad media de recaptura fue de 5.9 aftos (EE = 0.37; rango = 3-15 aftos). La edad media de adultos marcados 1993 o 1994 fue de 8.2 aftos (EE = 0.52) con un rango de 4 a 15 aftos. Las distancias dedispersi6n etre el sitio natal y el posterior sitio reproductivo estaban en el rango de 0 a 18.1 km (media = 8.2 km; EE --- 3.1) y no fue diferente a una distancia azarosa entre territorios.
[Traducci6n de Ivan Lazo]
Declines in numbers of Swainson's hawks (Buteo
swainsoni)n California (Bloom 1980, Risebrough
et al. 1989), Nevada (Oakleaf and Lucas 1976, Herron and Lucas 1978) and Oregon (Littlefield et al. 1984) have stimulated concern over long~term viability of this species' populations in the western United States. Numerous factors have been suggested as causes of regional declines, including loss of nesting habitat (Schlorff and Bloom 1983), conversion of foraging habitat to agriculture (Bloom 1980, Schmutz 1984, Estep 1989), livestock grazing
(Littlefield et al. 1984, Woodbridge 1991 ), predation
and interspecific competition (Littlefield et al. 1984, Janes 1987) and environmental contaminants (Henny and Kaiser 1979, Risebrough et al. 1989). Increased mortality during the nonbreeding season, when the bulk of the North American population
migrates to South America, has also been suggested
as a potential cause of noted declines (Bloom 1980,
White et al. 1989).
Neither causes nor remedies for population declines can be addressed without broad understanding of species biology and assessment of long-term patterns in population dynamics. Estimation of demographic parameters allows for analysis of population viability and trends, effects of management actions,
187
188 WOODBRIDGE ET AL. VOL. 29, NO. 3
identification of critical life history stages, and comparisons with other populations (Noon et al. 1992).
Such baseline data have proven essential in conservation of threatened and endangered species (Thomas et al. 1990, Noon et al. 1992).
We describe patterns of territory density, occupancy, productivity, age structure and dispersal in a population of Swainson's hawks over an 11-yr
period, and discuss variability in these estimates.
This study is part of a long-term investigation of the
ecology and demography of Swainson's hawks in
northern California.
STUDY AREA
This study took place in the Butte Valley in northern
California, approximately 10 km south of the Oregon
border and the Klamath Basin. The Butte Valley is included in the Modoc Plateau region of the California sagebrush steppe ecological province (Barbour and Major 1977) and is part of the basin and range physiographic
province (Franklin and Dyrness 1973). The study focused primarily on the unforested floor of the Butte Valley,
which ranges from 1280-1340 m in elevation and is approximately 415 km 2 in extent. Topography, soils, and natural vegetation of the Butte Valley are typical of the basin and range physiographic province (Franklin and
Dyrness 1973). Dominant vegetative associations are western juniper (Juniperus occidentalis) woodland (22%), grazed sagebrush steppe (21%), wetlands and seasonally flooded areas (5%), and agricultural fields (53%). Cultivated crops include irrigated alfalfa (Medicago satira), grains, and potatoes.
METHODS
We estimated the size and distribution of the breeding
Swainson's hawk population in the Butte Valley by conducting systematic surveys of the valley floor each year from 1983-94. Surveys in the adjacent Klamath Basin
consisted of visits to previously known territories. Open, flat terrain and high road densities in agricultural areas permitted good visibility and survey coverage from vehicles. Sage-steppe and juniper woodland habitats were surveyed on foot and from vehicles. In addition to searching for nest sites, we observed any foraging hawk of unknown origin until it returned to a nest site. Survey effort was concentrated during courtship and nest building (midApril through late May) although new nests were located as late as August.
We defined a territory as an area containing an active
nest and defended by a single pair of Swainson's hawks
in at least one year. In subsequent years, territories were classified as occupied if at least one adult was observed repeatedly during the early nesting season. A pair was
considered to have attempted to breed if a nest was constructed or new material added to an existing nest. Deserted nests were included in the sample if they showed signs of use in the present year (Postupalsky 1974, Steenhof and Kochert 1982). The number of nestlings and unhatched or broken eggs was recorded during at least two nest checks, the first occurring about 1 wk after the mean hatching date (16 June) and the second at banding (4 July to 15 August). We considered a breeding attempt to be
successful if >-1 young survived to 5 wk of age. Pairs that built nests but did not breed were included with failed attempts.
To control bias associated with survey timing, lower
detectability of failed breeding attempts (Steenhof and Kocheft 1982), and to account for permanently abandoned nests discovered in early years of the study, we included only territories which had been occupied at least once
during the four previous years in calculations of annual occupancy, nest success, and productivity. The exclusion of newly discovered territories from the annual calculations of annual suceess and productivity may result in a bias toward traditionally occupied territories with older breeders.
We banded nestling Swainson's hawks in the Butte
Valley each year from 1980-94. We also marked breeding
adults at the nest site by using a mist net with a live great horned owl (Bubo uirginianus) as a decoy (Hamerstrom
1963, Bloom 1987). Adult hawks were marked with individually numbered plastic legbands and U.S. Fish and Wildlife Service lock-on aluminum bands. Searches for
marked individuals were made each year at all known
occupied territories, as well as along established transects in foraging habitats. Mean values in the text are presented with standard errors.
RESULTS AND DISCUSSION
Population Size and Density. Swainson's hawks
occupying the Butte Valley during the breeding season were almost exclusively territorial adults; we recorded few nonterritorial adult floaters or subadults during the study. The high frequency of marked individuals (80% of breeders in 1994) in the population enabled us to distinguish nonbreeding floaters from territorial pairs and nonterritorial failed breeders.
The number of Swainson's hawk territories monitored during this study increased from 12 in 1984 to 83 in 1994. (Not all of these are included in
calculations of reproductive parameters.) Increases
during the earlier years of the study were due to
improved survey coverage; after 1990, however} fluctuations were related to colonization of new sites and abandonment of traditional territories. Since 1990,
14 new territories were established within the study
area, and four traditional territories were abandoned. The dependence of Butte Valley Swainson's hawks on limited high-quality foraging habitats such
as sprinkler-irrigated alfalfa was expressed in the
locations of newly colonized territories. Twelve of
the 14 new territories were established near fields
that had recently been converted to alfalfa cultivation.
SEPTEMBER 1995 SWAINSON'S HAWK POPULATIONS 189
Table 1. Annual occupancy, nest success, and fiedging
rate for Swainson's hawk territories in the Butte Valley, California, 1984-94.
YEAR
N PERCENT
TERRI- PERCENT SUCCESSTORIES a OCCUPIED FUL
MEAN NUMBER
OF YOUNG
PER NEST
ATTEMPT (SE)
1984 12 66.7 1.25 (0.37)
1985 13 77 60.0 1.30 (0.40)
1986 22 91 80.0 2.00 (0.26)
1987 34 88 73.3 1.71 (0.23)
1988 38 84 71.9 1.80 (0.20)
1989 45 84 65.8 1.30 (0.20)
1990 56 84 78.7 2.00 (0.20)
1991 64 86 52.7 1.50 (0.20)
1992 66 73 47.9 0.79 (0.14)
1993 64 61 51.3 1.10 (0.15)
1994 73 74 72.2 2.20 (0.15)
Mean 80.2 65.5 1.53
SE 2.8 3.4 0.14
a Number of territories sampled.
Most nests were located in western junipers, although four nests were constructed in ponderosa pine (Pinus ponderosa), one in an elm (Ulmus sp.), two
in basin bigsage (Artemisia tridentata). Territory distribution was strongly affected by the availability of patches of western juniper in close proximity to agricultural habitats used for foraging (Woodbridge 1991).
Overall territory density was 20/100 km 2. Territory density varied among five large (50-150 km 2) landscape blocks, ranging from 5.7/100 km 2 in irrigated pasture to 36.8/100 km 2 in a landscape dominated by alfalfa cultivation.
Territory Occupancy. The proportion of territories occupied in a given year ranged from 91% in 1986 to 61% in 1993, averaging 80% (SE -- 2.8)
over all years (Table 1). Fluctuations in annual occupancy rates were caused by non-use of traditional territories and by new pairs attempting to establish
territories in previously unoccupied areas. Occupancy also varied among individual territories, ranging from 14-90% of years monitored. At 52 territories with more than five consecutive years of monitoring data, we identified 35 traditional territories that were occupied during most (> 50%) years of the
study. Seventeen ephemeral territories exhibited low
(-• 50%) occupancy rates.
Reproductive Success. Between 1984 and 1994
we observed 454 nest attempts. Overall, pairs were
successful in 65% (SE = 3.6) of reproductive attempts, with annual success ranging from 48-80% (Table 1). Success rates were significantly lower
during the period from 1991-93. Nesting failure
between 1986 and 1990 typically resulted from loss
of nestlings (51% of failures), failure to lay eggs
(24%), failed incubation (17%) and incomplete nestbuilding (8%; Woodbridge 1991). Known nestling losses that resulted in nest failure were caused by
starvation (N = 21), predation by great horned owls
(N = 5) or golden eagles (Aquila chrysaetos) (N =
2), windstorms (N -- 5) and human disturbance (N
-- 3).
Fledging Rate. The mean fiedging rate for all
nest attempts was 1.53 (SE = 0.14) young per nest
attempt, ranging from 0.79 (SE-- 0.14) in 1992 to
2.20 (SE -- 0.15) in 1994 (Table 1). Mean annual
fledging rates varied among individual territories.
Of 52 territories with more than five consecutive
years of monitoring data, 11% exhibited low (<0.5
young/attempt) fledging rates; an additional 11%
fledged <1.0 young/attempt. Fledging rates were
> 1.50 young/attempt at 42% of monitored territories. Territories with consistently low reproduction were associated with sage-steppe habitats, whereas
agricultural habitats supported higher reproductive
rates (Woodbridge 1991).
Mark-recapture Results. We banded 567 nestling Swainson's hawks in the Butte Valley between 1979 and 1994. Forty-one of these marked nestlings
were recaptured as breeding adults in subsequent
years, and nine were recovered as post-fledging mortalities within the Butte Valley. An additional eight were recovered in Latin America, giving an overall
band return rate of 10.2%. The virtual absence of
subadult Swainson's hawks in the study area, however, suggests that these birds spend their second summer elsewhere, and may not be available for
recapture. This would result in underestimation of
band return rates, which are largely based on recapture of breeding adults.
Population Age Structure. Swainson's hawks
occupying the Butte Valley during the breeding season were almost exclusively in adult plumage; we recorded only 12 individuals in subadult plumage.
Subadults observed early in the season (April to
May) were typically absent after mid-May, and were
assumed to be late migrants. Paired territorial
Swainsoffs hawks in subadult plumage were re-
190 WOODBRIDGE ET AL. VOL. 29, NO. 3
9 6
O• 8
.--
;4 •o 3
0 --• 0
3 4 5 6 7 8 9 10 11 12 13 14 15 3 4 5 6 7 8 9 10 11 12 13 14 15 Age in Years Age in Years
Figure 1. Frequency distribution of age at recapture of 41 Swainson's hawks banded as nestlings in the Butte
Valley, California, 1980-94.
Figure 2. Frequency distribution of ages for 36 Swainson's hawks banded as nestlings in the Butte Valley, California (1980-92) and observed in 1993-94. corded on four occasions; in all cases the subadult
hawk was female. One subadult male acted as a nest
helper, provisioning young at a nest also tended by
an adult pair.
Forty-one Swainson's hawks banded as nestlings
•n the Butte Valley have been recaptured as breeders
since 1981. The mean age at recapture was 5.9 yr
(SE = 0.37), ranging from 3-12 yr (Fig. 1). Age at
recapture should not be interpreted as age of first
breeding; many of these individuals may have bred
elsewhere or eluded capture for several years before
recapture. Of five 3-yr-old Swainson's hawks recaptured as territorial breeders (assumed first breeding), two were still in subadult plumage, suggesting some variability in maturation rates.
Approximately 80% of the 41 recaptured and color-marked Swainson's hawks were observed at least once in subsequent years. The age distribution of
color-marked birds observed in either 1993 or 1994
(N = 36) ranged from 4-15 yr, with a mean age of
8.2 yr (SE = 0.52; Fig. 2). We believe that this is
our best estimate of the age distribution of the breeding population. Mortalities of known-age Swainson's hawks were caused by collision with a train (one 6-yr-old male) and probable collision with vehicle (one 15-yr-old male).
Dispersal. We used natal dispersal distances and
interterritory movements of marked adult Swainson's hawks to assess immigration and emigration in this study area. For 41 nestlings recaptured as
breeding adults (20 female, 21 male), dispersal distances ranged from 0-18.1 km, and averaged 8.8 km (SE = 1.1 km). Mean distance between natal and
subsequent breeding sites was not statistically different from random interterritory distance (t = 1.06, P = 0.17, two-tailed ), and did not differ between
sexes (t = 1.33, P = 0.12, two-tailed ). Our estimate
of dispersal distance may be biased low, since survey
effort is not as extensive in the adjacent Klamath
Basin. Two breeding adults recaptured in the Butte
Valley had moved 10.2 km and 7.9 km from natal
sites in the adjacent Klamath Basin. We recorded
only one confirmed case of natal dispersal to a breeding territory outside of the Butte Valley; this bird was found breeding at a territory in the Klamath
Basin, 36.8 km from its natal site. A nestling banded
in 1982 was found injured in April 1988 in Diamond
Valley, Nevada, 565 km southeast from its natal site.
Another nestling banded in 1979 was found injured
in August 1981, in Christmas Valley, Oregon, 160
km northeast of its natal site. Whether these hawks
were breeding at their recovery sites is unknown.
Dispersal of hawks into the Butte Valley from
natal sites in the Klamath Basin suggests that the
Butte Valley is not a closed population, and there
may be substantial genetic interchange with the
Klamath Basin area. The two areas are likely interacting elements of one metapopulation. Territory occupancy and reproductive success, however, differed dramatically between the two areas during the same period. Bloom and Hawks (in Risebrough et
al. 1989) reported <50% territory occupancy and
< 50% nest success in the Klamath Basin in the mid1980s, considerably less than in the Butte Valley.
Because of unequal levels of marking and monitoring
in the two areas, we were unable to quantify the
level of exchange.
Within the Butte Valley study area, we recorded
25 interterritory movements by marked adult Swainson's hawks (11 male, 14 female). These movements ranged from 0.97-6.3 km (mean -- 2.2, SE = 0.23
km), and typically were short moves between neigh-
SEPTEMBER 1995 SWAINSON'S HAWK POPULATIONS 191
boring territories. Mean adult dispersal distance was
significantly less than the mean nearest-neighbor
distance (3.7 km, SE -- 0.87) for the Butte Valley
population (t = 6.77, P = 0.03, one-tailed). Of 36
adult hawks marked in the adjacent Klamath Basin
between 1981 and 1988, none were observed within
the Butte Valley.
While evidence of natal dispersal and individual
movements to and from our study area are valuable
preliminary data, assessments of immigration and
emigration would be greatly enhanced by more comprehensive monitoring of breeding Swainson's hawks in the neighboring Klamath Basin. Among other
things, this would help us assess the extent to which
the Butte Valley serves as a source population for
more marginal habitat areas to the north and east.
Concluding Remarks. Because Swainson's hawks
are long-lived, description of population demography requires long-term study of marked populations in order to account for the effects of generation time
and environmental variability (Newton 1979, Noon
et al. 1992). Temporal variability in reproductive
performance such as we have observed during the
course of this study may or may not translate into
long-term changes in the local Swainson's hawk population. Assessment of critical values may best be accomplished through mathematical modelling of
population trends based on life history matrices.
Identification of conservation measures for maintaining the long-term viability of this population depends on analysis of the extent to which differences
in reproductive success and nestling survival are habitat related. The presence of both traditional and marginal ephemeral territories in our study area
suggested that a subset of territories contributed disproportionately to the long-term viability of the population. In our area, cultivated alfalfa appears to have become a critical habitat element (see Woodbridge 1991), replacing the productive native grasslands which were the original vegetation. Conservation measures that restore productive perennial grasslands and open shrubsteppe habitats on public
lands and Conservation Reserve lands in the Butte
Valley might ensure the resilience of Swainson's
hawks to changes in agricultural ecomomics and
practices.
ACKNOWLEDGMENTS
We owe thanks to the many field technicians and student interns who contributed to this project over the years. C.
Cheyne, S. Baker, J. Montejo Diaz, A. Strassler and T.
Seaget in particular were instumental to the success of this project. T. Farmer and J. Stout of the Goosenest
Ranger District, Klamath National Forest, provided financial and logistical support for our studies in the Butte Valley. The assistance and camaraderie of S. Hawks, B.
Clark, E. and J. Henckel and S. Bales in collection of
Klamath Basin data is gratefully acknowledged. Partial
funding was provided by Brian Walton of the Predatory
Bird Research Group.
LITERATURE CITED
BARBOUR, M.G. AND J. MAJOR. 1977. Terrestrial vegetation of California. John Wiley and Sons, New York, NY U.S.A.
BLOOM, P.H. 1980. The status of the Swainson's hawk
in California, 1979. Final Report II-8.0, USDI Bur.
Land Manage. and Fed. Aid in Restoration, Proj. W-54-
R-12, Calif. Dept. Fish and Game, Sacramento, CA
U.S.A.
1987. Capturing and handling raptors. Pages
99-123 in B.A. Gerin Pendleton, B.A. Millsap, K.W
Cline and D.A. Bird lEDs. I, Raptor management techniques manual. Natl. Wildl. Fed., Washington, DC U.S.A.
ESTEP, J.A. 1989. Biology, movements, and habitat relationships of the Swainson's hawk in the Central Valley of California, 1986-1987. Calif. Dept. Fish and Game, Nongame Bird and Mammal Sec. Rep. Sacramento, CA U.S.A.
FRANKLIN, J.F. AND C.T. DYRNESS. 1973. Natural vegetation of Oregon and Washington. USDA For. Serv.
Gen. Tech. Rep. PNW-8. Corvallis, OR U.S.A.
HAMERSTROM, F. 1963. The use of great horned owls
in catching marsh hawks. Proc. Internat. Ornithol. Congr 13:866-869.
BENNY, C.J. AND T.E. KAISER. 1979. Organochlorine
and mercury residues in Swainson's hawk eggs from
the Pacific Northwest. Murrelet 60:2-5.
HERRON, G.B. AND P.B. LUCAS. 1978. Population surveys, species distribution, and key habitats of selected nongame species. Nev. Dept. Fish and Game, Job Perform. Rep., Proj. W-43-R, Study 1, Jobs 1 and 2, Reno, NV U.S.A.
JANES, S.W. 1987. Status and decline of Swainson's
hawks in Oregon: the role of habitat and interspecific
competition. Oreg. Birds 13:165-179.
LITTLEFIELD, C.D., S.P. THOMPSON AND B.D. EHLERS.
1984. History and present status of Swainson's hawks
in southeast Oregon. Raptor Res. 18:1-5.
NEWTON, I. 1979. Population ecology of raptors. Buteo
Books, Vermillion, SD U.S.A.
NOON, B.R., K.S. MCKELVEY, D.W. LUTZ, W.S. LAHAYE,
R.J. GUTIERREZ AND C.A. MOEN. 1992. Estimates
of demographic parameters and rates of population
change. Pages 175-186 in J. Verner, K.S. McKelvey,
B.R. Noon, R.J. Gutierrez, G.I. Gould and T.W. Beck
[TECH. EDS.], The California spotted owl: a technical
192 WOODBRIDGE ET AL. VOL. 29, NO. 3
assessment of its current status. Gen. Tech. Rep. PSWGTR-133. USDA For. Serv., Albany, CA U.S.A.
OAKLV, AF, R.J. AND P. LUCAS. 1976. Population surveys, species distribution, and key habitats of selected nongame species. Nev. Dept. Fish and Game, Job Perform.
Rep., Proj. W-53-R, Study 1, Jobs 1 and 2, Reno, NV
U.S.A.
POSTUPALSKY, S. 1974. Raptor reproductive success: some problems with methods, criteria and terminology. Pages
21-31 in F.N. Hamerstrom, Jr., B.E. Harrell and
R.R. Olendorff lEDS.], Management of raptors. Raptor
Res. Found., Vermillion, SD U.S.A.
RISEBROUGH, R.W., R.W. SCHLORFF, P.H. BLOOM AND
E.E. LITTRI•LL. 1989. Investigations of the decline
of Swainson's hawk populations in California. J. Raptor Res. 23:63-71.
SCHLORFF, R.W. AND P.H. BLOOM. 1983. Importance
of riparian systems to nesting Swainson's hawks in the
Central Valley of California. Pages 612-618 in R.E.
Warner and K.M. Hendrix [EDS.], California riparian
systems. Univ. California Press, Berkeley, CA U.S.A.
SCHMUTZ, J.K. 1984. Ferruginous and Swainson's hawk
abundance in relation to land use in southeastern Alberta. J. Wildl. Manage. 48:1180-1187.
STEENHOF, m. AND M.N. KOCHERT. 1982. An evaluation of techniques used to estimate raptor nesting success. J. Wildl. Manage. 46:885-893.
THOMAS, J.W., E.D. FORSMAN, J.B. LINT, E.C. MY, SLOW,
B.R. NOON AND J. VERNER. 1990. A conservation
strategy for the northern spotted owl. U.S. Government
Printing Office, Washington, DC U.S.A.
WHITE, c.m., D.A. BOYCE AND R. STRANEK. 1989. Observations on Buteo swainsoni in Argentina, 1984. Pages 79-87 in B.-U. Meyburg and R.D. Chancellor [EDS.],
Raptors in the modern world. World Working Group
on Birds of Prey and Owls, Berlin, Germany.
WOODBRIDGE, B. 1991. Habitat selection by nesting
Swainson's hawks: a hierarchical approach. M.S. thesis, Oregon State Univ., Corvallis, OR U.S.A.
Received 20 January 1995; accepted 22 May 1995
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OPINION ARTICLE
Using performance standards to guide vernal pool
restoration and adaptive management
Karen J. Schlatter1, Akasha M. Faist2, Sharon K. Collinge3 Although many restoration projects now include monitoring and evaluation in an adaptive management approach, a failure to employ distinct performance standards can lead to inconsistent and unclear results that may hinder learning from project outcomes and complicate large-scale assessments of restoration success. Such is the case with vernal pool restoration projects in California, where performance standard guidelines are vague and inconsistently applied across agencies implementing restoration projects. However, positive steps have been made in recent years to develop wetland functional assessments and monitoring protocols in California to reduce inconsistencies and promote ecologically meaningful restoration. Additional work is needed to develop specific guidelines for vernal pool restoration performance standards and define their role within an adaptive management framework. We provide a case study of a vernal pool restoration project in central California to illustrate some of the challenges in using currently available vernal pool performance standard guidelines and propose suggestions for increasing their ecological relevance and clarity.
Key words: California, evaluation, functional assessment, mitigation, monitoring, success criteria, wetland Implications for Practice
• Current vernal pool restoration performance standards are often vague, challenging to access, inconsistent across agencies implementing projects, and lack ecological relevance. • The absence of well-defined and ecologically meaningful performance standards is hindering adaptive management
and large-scale assessments of restoration success.
• Regional performance standard guidelines should be
developed that incorporate the following: assessments of performance variation (not just average performance);
specific and repeatable measurement methods for performance standard metrics; at least a 5-year monitoring timeframe to assess performance; and greater emphasis
on assessments of ecosystem function over form.
Introduction
A number of studies have illustrated the need to include measurements of “success” to evaluate if and to what extent objectives have been met in ecological restoration and habitat creation projects (Hobbs & Harris 2001; Palmer et al. 2005; Bernhardt et al. 2007; Suding 2011). Despite this observed need, the incorporation of empirical evaluations of restoration success into management plans has generally lagged behind the research (Wortley et al. 2013). Without well-defined and ecologically meaningful evaluation standards, a critical step is missing in the process of adaptive management to achieve high levels of restoration success (Palmer et al. 2005; Suding 2011). Although the number of academic publications demonstrating empirical evaluations of restoration projects has increased in recent years (Wortley et al. 2013), it is questionable how these evaluation methods are being incorporated into a majority of restoration projects with limited financial resources and focus on research.
Monitoring and adaptive management are often required for many government directed restoration and mitigation projects (USFWS 2005; USACOE & USEPA 2008; Williams 2011). The
push for adaptive management suggests a trend toward using monitoring as a way to “learn by doing”—employing an experimental design to test management strategies and incorporating results into management decisions (Williams 2011). Critical to this process is the establishment of quantitative measurements of success, or performance standards, which are measureable metrics used to assess if the project meets its objectives (Harman et al. 2012). Performance standards provide a direct link between project goals and outcomes, enabling practitioners to determine the effectiveness of the project quantitatively and assess best practices and approaches. Performance standards are embedded in all steps of the adaptive management cycle; however, they are often unclear or deemphasized in the project planning process, with little guidance on how they should be Author contributions: SC designed, managed, monitored, and evaluated the vernal pool restoration project and initially reviewed performance standards used in the project; KS conducted research on vernal pool performance standards and wrote and edited the manuscript; AF, SC contributed ideas and revisions to the manuscript. 1University of Colorado Environmental Studies Program, UCB 397, Boulder, CO 80309, U.S.A.
2Department of Ecology and Evolutionary Biology, University of Colorado, UCB 334, Boulder, CO 80309, U.S.A.
3Address correspondence to S. K. Collinge, email [email protected] © 2016 Society for Ecological Restoration
doi: 10.1111/rec.12326
March 2016 Restoration Ecology Vol. 24, No. 2, pp. 145–152 145 Performance standards
determined and the role they play in adaptive management (Bernhardt et al. 2007). Often, information gathered from ecological monitoring is not related to project outcomes, nor used to determine remedial actions (Palmer & Hondula 2014).
Defining objective and measurable performance standards can be challenging for restoration practitioners due to the lack of easily accessible scientific information and overarching regional management goals, differences in stakeholder values, and the uncertainty posed by rapidly shifting baseline conditions (Suding 2011). Yet, habitat restoration and mitigation projects are likely to be either misinformed or difficult to conduct in the absence of broadly accepted performance standards. Such is the case with vernal pool habitat creation and restoration in California. Millions of dollars are spent every year on wetland restoration projects in California alone (CWMW 2010); yet such expenditures can seem ineffective if the state cannot report on the success of such projects, provide a valid assessment of overall ecosystem health, or effectively adaptively manage systems using feedback from project evaluations.
Restoration efforts on stream ecosystems have experienced similar growing pains in identifying meaningful measures of restoration success over the past 10 years (Palmer et al. 2005; Bernhardt et al. 2007; Woolsey et al. 2007; Doyle & Shields 2012; Palmer & Hondula 2014). Although we focus specifically on vernal pool systems in this article, we raise issues of performance standards that apply to a broad range of restoration projects in varying ecosystems and locations. We assess success of our vernal pool mitigation project in central California, synthesizing lessons learned from our long-term monitoring data set with additional research on performance standards to build upon recommendations for measuring success. The development of a revised list of vernal pool performance standards and measurement methods is outside the scope of this article; but, by drawing attention to the prevalent issues of inconsistency and ambiguity in existing vernal pool performance standards, we aim to stimulate discussion, research, and eventually progress toward the development of well-defined and ecologically relevant standards.
Vernal Pool Ecology and Restoration Efforts
Primarily distributed in regions with Mediterranean climates and some non-Mediterranean climates including temperate forest systems (Brooks & Hayashi 2002), vernal pools are ephemeral wetlands that experience inundation from winter rains and desiccation in the summer and fall (Keeley & Zedler 1998). Vernal pools are critical areas of floral and faunal diversity and typically support a high number of endemic species due to their unique and extreme seasonal conditions (Keeley & Zedler 1998). However, land-use change has resulted in the loss of 60–95% of vernal pool habitat in California alone (Barbour et al. 2005, 2007). Many vernal pool species are classified as endangered or threatened (USFWS 2005), which has prompted the establishment of vernal pool recovery programs and restoration efforts.
Vernal pool restoration was first attempted in the 1980s (Zedler 1987), and the number of projects has increased in the past 30 years. However, the unique hydrological regime of vernal pools can be difficult to reestablish, which often results in unsuccessful restoration attempts (De Weese 1998; Calhoun et al. 2014). The difficulty in creating and restoring vernal pool habitat is particularly an issue for mitigation projects that allow for the destruction of intact vernal pools on the premise that constructed vernal pools will replace natural pools’ ecological function and structure (Black & Zedler 1998; Calhoun et al. 2014). A review of wetland construction projects in California found that constructed vernal pools received the lowest ratings for replacing lost habitat values (De Weese 1998). Calhoun et al. (2014) suggest that vernal pool construction projects should be used as a “last resort” only when attempts at protection have been exhausted.
In light of such challenges in vernal pool creation and restoration projects, well-defined measurements of success and repeatable evaluation methods would be valuable to managers in assessing why failures occur and how projects could be better implemented in the future. In particular, project performance standards developed from reference pool sites are critical in providing quantitative project goals that can stipulate further management actions if they are not met.
The Evolution of Performance Standards
Although significant progress has been made in attempts to establish meaningful ecological assessments, government initiatives outlining performance standard guidelines for wetland and stream restoration projects demonstrate that the evolution is not complete. The U.S. Army Corps of Engineers (USACOE) and U.S. Environmental Protection Agency (USEPA) provided guidance in establishing wetland systems performance standards in the 2008 Compensatory Mitigation for Losses of Aquatic Resources Rule (USACOE & USEPA 2008). The document’s vague language, use of form-based performance standards over function-based standards, and lack of integration of standards into adaptive management have resulted in regional inconsistencies, unsuccessful restoration of ecosystem function, and project permits that lack any specified performance standards (Bronner et al. 2013). Similarly, the “Function-Based Framework for Stream Assessment and Restoration Projects” by the USEPA (Harman et al. 2012) attempted to provide guidance on defining performance standards and function-based assessments for stream restoration projects; however, the document lacks operationalization, and several suggested metrics do not actually measure ecosystem function (Palmer et al. 2014).
Looking more specifically at vernal pool performance standards, a review of California vernal pool mitigation projects in 1998 found that the standards in use were “insufficient to assure successful habitat replacement,” with high variability in the types of parameters evaluated across agencies and projects (De Weese 1998:217). A more recent study on the classification of vernal pools in California determined that performance standards were lacking standardization as to what plant community attributes are measured, had variable lengths of monitoring time following project implementation, and metrics intended to 146 Restoration Ecology March 2016
Performance standards
evaluate ecological parameters were not ecologically relevant (Barbour et al. 2007). A 2014 review of vernal pool creation projects in the northeastern United States provides a similar critique of performance standards, noting that they are insufficient indicators of restored habitat function, particularly with regard to the maintenance of faunal communities (Calhoun et al. 2014).
Collaborative groups such as the California Wetland Monitoring Workgroup (CWMW) have recognized the issue of dispersed responsibility among agencies implementing wetland restoration projects in California (CWMW 2010) and developed the California Rapid Assessment Method (CRAM) in 2009 as a way to address the lack of standardized monitoring protocols (CWMW 2010). Although the use of CRAM is not currently mandated for agencies implementing wetland restoration projects, it is an accessible and useful resource for managers who wish to use it. CRAM is highly encouraged by regulatory agencies and is already being adopted in many projects (C.
Schaefer 2014, Schaefer Ecological Solutions, personal communication).
Additionally, Bauder et al. (2009) provide useful guidelines for using a hydrogeomorphic (HGM) approach to assess wetland function of vernal pools in Southern California. Their approach includes four key components: “(a) the HGM classification, (b) identification of reference wetlands, (c) assessment models/functional indices and (d) assessment protocols” (Bauder et al. 2009:4). Particularly of use to vernal pool restoration practitioners, they provide rigorous scientific justification for parameters that should be used to indicate vernal pool function as well as detailed monitoring protocols for various metrics (Bauder et al. 2009). Bauder et al. (2009) proposed the following as critical functions specific to vernal pools in Southern California: (1) surface and sub-surface water storage, (2) hydrologic networks, (3) biogeochemical processes, (4) maintenance of characteristic plant community, and (5) maintenance of characteristic faunal community.
Bauder et al.’s (2009) approach represents an increasing emphasis on functional assessments and process-based restoration over form-based restoration and evaluation (see Doyle & Shields 2012; Bronner et al. 2013; Palmer et al. 2014). Evaluations of ecosystem function can include both assessments of dynamic processes and maintenance of structural components needed for ecosystem functioning, as is exemplified in Bauder et al.’s (2009) inventory of vernal pool functions. Functional assessments are becoming increasingly popular in the field of restoration ecology, as noted in the performance standard critiques listed above. While we should strive to incorporate such assessments into restoration projects, they typically require more time and money, and additional research is needed to determine their applicability in the field and their proven effectiveness in predicting ecosystem function (Palmer et al. 2014).
Although some agencies have incorporated various recommendations of the above sources as well as metrics from the HGM guidelines (Bauder et al. 2009) into restoration and habitat conservation plans (HCP) (see Solano County HCP and Draft San Diego Vernal Pool HCP, City of San Diego 2015), the majority of performance standards are still determined on a project-by-project basis, with little to no overarching guidelines or requirements as to what they should be. Ultimately, it is up to the agency and even the specific project manager to determine monitoring methods, performance standards, and any regulative actions if a project is deemed unsuccessful. Thus, while we recognize that the HGM and CRAM wetland assessment approaches as well as the extensive research supporting the proposed methods are critical and necessary steps in the development of appropriate performance standards, we believe that regional guidelines are specifically needed for the development of vernal pool performance standards, including guidelines on the integration of standards into adaptive management plans.
In the following section, we provide a case study in which we applied vernal pool performance standards using a long-term data set from our vernal pool flora mitigation project in central California. The purpose of the case study is to demonstrate issues we encountered in the application of performance standards to our project and provide fundamental insights based on our experience. We first assess our project results through performance standards defined by De Weese (1998), provide a comparison with more recent standards developed by Barbour et al. (2007), and finally, list recommendations for further improvements on these standards. While we focus primarily on vernal pool floral community assessments, our aim is that the observations can be applied more broadly to faunal community assessments and the development of performance standards in general. We do not provide detailed data analyses in the case study (see previous studies by Collinge) or focus on suggested detailed monitoring methods, as that is beyond the scope of this opinion piece.
Case Study
In 1999, we implemented a vernal pool creation project in a 15-ha study site that historically supported a vernal pool complex in Solano County, California, to mitigate for the endangered plant species, Lasthenia conjugens (Collinge et al. 2013).
Eighty naturally occurring pools present at the site were used as references to guide construction of 256 pools and to measure success of constructed pools (Collinge et al. 2013).
We used performance standards outlined in De Weese (1998), which were based on USFWS 1994 draft vernal pool mitigation and monitoring guidelines, to inform our monitoring efforts and determine the metrics used to evaluate project success, as this was one of the few resources available on performance standards at the time the pools were constructed. We monitored hydrological and vegetation parameters of constructed pools and a subset of reference pools beginning in 1999 (see Collinge et al. 2013 for detailed monitoring methods) and conducted a project evaluation in 2011 based on De Weese (1998) (see Table 1).
Hydrological Monitoring
Hydrology performance standards (De Weese 1998) specify that: (1) the depth of inundation of constructed pools should March 2016 Restoration Ecology 147
Performance standards
Table 1. Comparison of published vernal pool performance standards. (VPE stands for vernal pool endemic.) Performance Standard Source
Hydrology De Weese 1998 Barbour et al. 2007 Solano County HCP 2012 Pool depth Maximum depth of inundation is
within range of reference pools.
Depth and/or duration of ponded
water in created pools should not
differ statistically (p = 0.05) from
those in nearby natural pools.
No hydrology performance
criteria in vernal pool
section. Hydrology and
water quality monitoring to
be determined/conducted
by individual reserves
according to resource
management plan.
Duration of inundation Longest period of inundation not greater than 125% of reference
pools.
Vegetation
Absolute and relative cover Absolute and relative cover by VPEs in constructed pool should be no
less than the minimum recorded
in reference pools.
Absolute and relative cover by
vernal pool endemics in
constructed pools should not be
statistically different (p = 0.05)
from the average values in ref
pools.
Absolute and relative cover of
each vernal pool endemic
in restored pools shall be
statistically similar to the
range of values of each
species found in reference
pools.
Species richness Each constructed pool must support
no fewer than the lowest number
of VPEs recorded in reference
pools.
The number of vernal pool endemics
in constructed pools should not be
statistically lower (p = 0.05) than
the average number of those taxa
among reference pools.
The number of vernal pool
endemics in restored pools
shall be statistically similar
to the average number of
those taxa among reference
pools.
Reproductive vigor VPEs shared by both the impact and
reference pools shall be as
vigorous and reproductively
active in the constructed pools as
the reference pools.
The vigor (biomass accumulation)
and reproductive activity (seed
production) of VPEs in
constructed pools should not be
statistically lower (p = 0.05) than
those of the same species in
reference pools.
Criterion deleted.
Dominant species By last year of monitoring, any
VPEs that are dominant (relative
cover of at least 20%) in at least
30% of reference pools shall be
present as a dominant species in
all constructed pools.
Recommend deletion. Criterion deleted.
Community types None The identity of community types in created pools and the mixture in
which they occur should match
that of reference pools (using a
Sorensen Similarity Index
formula where “matching” means
an SSI > 50%).
There shall be no significant
difference in the ratio of
restored and reference plots
observed versus the ratio of
restored and reference
plots expected between the
first three groups formed in
the cluster analysis
corresponding to deep,
shallow, and edge
community types.
Reference pools should be chosen
subjectively so that collectively
they represent the diversity of
species and communities that
exist in the pools to be taken.
Nonnative species None The number and cover of exotic
species in any constructed pool
should not be significantly higher
than the average among reference
pools.
The number and cover of
nonnative species in any
restored pool shall be
statistically similar to or
lower than the average
among reference pools.
148 Restoration Ecology March 2016
Performance standards
be within the range of that observed in reference pools, and (2) that the longest period of inundation in constructed pools should not be greater than 125% of that found in reference pools. We found that one fourth to one third of our constructed pools did not meet the “depth of inundation” hydrological performance standard. Most constructed pools that did not fall within the range of reference pools were too shallow. Although the majority of pools met the second hydrology standard, 18–50% of constructed pools (depending on the monitoring year) did not meet the minimum period of inundation in reference pools, an important variable for vernal pool invertebrate species reproduction (Philippi et al. 2001). In addition, the “period of inundation” can be interpreted differently by the practitioner, that is, by including only the period of continuous inundation or by including any date on which a point in the pool was inundated, again highlighting issues of ambiguity that can lead to inconsistency in project evaluations. Although we used the second definition, continuity of inundation may be important for determining the success of native (especially vs. exotic) vernal pool species (Gerhardt & Collinge 2003; 2007).
While Barbour et al. (2007) attempt to provide some clarifications and improvements to the De Weese (1998) vernal pool hydrology performance standards (see Table 1), definitions and suggested metrics are still somewhat vague. They suggest that the “depth and/or duration of ponded water in created pools should not differ statistically (p = 0.05) from those in nearby natural pools” (Barbour et al. 2007:108). Although not specified, presumably it is the average depth and/or duration of ponded water of the constructed pools that should not statistically differ from the average of reference pools for a given year.
The use of averages in constructed and reference pool comparisons can be misleading in evaluations of success, as we discuss in greater detail in the “principles” section below. In addition, it would be helpful to describe specifically how depth and duration should be measured. Similarly, the standard for ponded water duration is vague, leaving the practitioner to define duration as continuous or the total sum of days inundated.
Vegetation Monitoring
Vegetation performance standards state that absolute cover, species richness, vigor and reproductive activity, and dominance by vernal pool endemics (VPEs) shall be similar in constructed and reference pools (see Table 1) (De Weese 1998). On average, many of our constructed pools did meet at least one of the vegetation performance standards on a yearly basis. In 2003, the average abundance of our target plant species, L. conjugens was not significantly different in constructed pools and reference pools (Collinge 2003). However, by 2009, one third of constructed pools contained little to no VPE species despite employing the same management regime for constructed and reference pools over the duration of the project (Collinge et al. 2013). This suggests that although initial restoration efforts may have been considered successful according to mitigation project performance standards, the long-term restoration outcomes were notably less positive.
We also found the De Weese (1998) standard of vigor and reproductive activity to be vague and difficult to determine.
We compared the number of viable seeds from the VPE L.
conjugens (Contra Costa goldfields) in constructed and reference pools as a measure of plant vigor, and also noted that the increased abundance of vernal pool plant species in constructed pools during the first few years of the experiment suggested seed viability.
Barbour et al. (2007) provide a series of useful suggestions for clarifying and improving the De Weese (1998) vegetation performance standards including measurable definitions of vigor (biomass accumulation) and reproductive activity (seed production), criterion to address exotic species, and several new criteria that are community oriented rather than species oriented.
Similarly, Calhoun et al. (2014) do not recommend the use of species richness as a performance standard because richness may reflect a high number of exotic species, and does not necessarily relate to vernal pool health. Rather, community composition is a suggested standard that can provide a more ecologically relevant measure of success (Calhoun et al. 2014). Lastly, Barbour et al. (2007) suggest deleting the De Weese (1998) criterion of VPE species dominance, as they contend it does not promote heterogeneity of constructed pools.
Principles for Improved Performance Standards
Our case study demonstrates the following issues with vernal pool performance standards (see Table 2): (1) using averages to evaluate performance can be misleading of true project outcomes; (2) performance standards can be vague and lack specific measurement methods; (3) metrics used may not be ecologically relevant or measure ecosystem function (e.g. species richness vs. community composition); (4) short-term monitoring may not indicate long-term success and/or sustainability of restoration goals. While there may be revised and more recent vernal pool performance standards that address some of these issues, they are not easily accessible as our research efforts on this topic proved. It was exceedingly difficult to find information on vernal pool performance standards currently in use in mitigation and restoration projects in California, even when contacting the agencies overseeing ongoing projects. Some recently developed HCPs do list specific performance standards, but they similarly embody some of the issues we outline above.
Synthesizing lessons learned from our experience and
research conducted by others, we developed the following key principles for improving vernal pool performance standards (see Table 2). We consider this a preliminary set of principles which researchers, agencies, and practitioners can debate and build upon to encourage the development of ecologically relevant, clear, consistent, and accessible performance standards.
First, an assessment of performance variation can provide a critical and more complete evaluation of project success in the case where “failed” constructed vernal pools are balanced out by successful constructed pools in the calculation of the average performance. Variance can also be used as an indicator of ecosystem regime shifts (Scheffer 2009), which could March 2016 Restoration Ecology 149
Performance standards
Table 2. Issues encountered in our vernal pool restoration project evaluation and suggested actions for improving performance standards.
Issues Encountered Suggested Actions
Inconsistent use of performance standards across agencies and projects
Standardize performance standards on a regional scale by creating standards based on comparisons of restored pools to regionally relevant, naturally occurring pools
Terms used in performance standards can be vague and lack specific measurement methods
Clear definitions and specific and repeatable methods for measuring metrics described in standards should be provided in performance standards guidelines
Using averages to evaluate performance can be misleading of true project outcomes
Assess variation in performance (potential use of a
predetermined failure rate metric)
Metrics used may not be ecologically relevant or measure ecosystem function
Performance standard metrics should ideally measure
ecosystem function rather than form
Short-term monitoring may not indicate long-term success and/or sustainability of restoration goals
Five-year minimum length of monitoring time, but a longer timeframe is preferred
Lack of easily accessible information on vernal pool
performance standards
Make standardized set of suggested regional performance standards available on agency and practitioner websites Monitoring results often not used in an adaptive
management process
HCPs and recovery plans should outline adaptive management steps including the use of monitoring data to reevaluate project goals and restoration practices and provide suggested actions if performance standards are not met
be helpful in evaluating both constructed and reference pool functional health and determining management strategies. A useful metric related to variance could be the number (and percentage) of constructed pools that had zero frequency of native species (i.e. all cover was by exotics). This statistic would show that although the average abundance of VPEs
did not differ significantly between constructed and reference pools, there were some “failed” constructed pools that did not meet the performance standard. A predetermined “failure rate” could provide a threshold to determine if additional restoration actions are needed to achieve a higher level of vernal pool creation success. In addition, a similarity index at the community level could be an informative metric to assess how similar restored vernal pool communities are to reference pools.
Second, our case study illustrates the potential gap in the adaptive management cycle in which monitoring results should be used to inform and revise management goals and practices.
Our monitoring data suggested that one third of the constructed pools “failed” due to shallow pool depth and accumulation of invasive plant species litter. A subset of failed pools were manually re-excavated and seeded to promote the establishment of our target plant species, although this was not mandated nor funded by the managing agency. HCPs and other recovery plans should outline steps in the adaptive management process including the use of monitoring data to reevaluate project goals and restoration practices and require additional management actions if performance standards are not met.
In addition, it is assumed but worthy to note that performance standard guidelines should be based on a comparison of restored pools to appropriate, naturally occurring reference pools that are regionally relevant (both within and outside of the site if possible) (Barbour et al. 2007; Bauder et al. 2009). To create appropriate performance standards, data collection from reference pools should take place over time rather than in one season to account for system dynamics (which are highly variable for vernal pools) and similarly, restored pool monitoring should take place over a different seasons (De Weese 1998; Bauder et al. 2009; Calhoun et al. 2014).
We propose that through the use of site-appropriate reference pool comparisons, performance standards could be standardized on a regional or statewide scale; for example, the identity of community types in created pools and the mixture in which they occur should match that of reference pools (taken from Barbour et al. 2007, see Table 1). While vernal pools distributed across the region will obviously vary in species composition, size, hydrology, and other variables, using reference pools to define performance standards essentially standardizes the assessment through the comparison of site-specific variables. In this manner, success can be measured even as overall conditions are changing, as opposed to assessing success based on a fixed idea of general vernal pool attributes. While we acknowledge that not all projects will have the same goals and thus the same standards, making such regional performance standard guidelines widely available (but not obligatory) to practitioners would greatly help to reduce inconsistencies across agencies and would enable assessments of restoration success and vernal pool health on a broader ecological scale.
Our case study highlights the focus on form versus
function-based restoration and monitoring; the mitigation project was designed to establish a specific plant species, and monitoring was focused on components that would indicate success of that goal. As a result of our project’s narrow focus on flora, the vernal pool faunal community was entirely ignored, even though the recovery of endemic vernal pool faunal populations is a goal for other restoration projects in the region and nation-wide. The plant community is not indicative of faunal community health, nor is it an accurate reflection of the physical, chemical, and biological processes that shape ecosystem function (Cole 2002). As the development of function-based 150 Restoration Ecology March 2016
Performance standards
assessments evolves, performance standard metrics should ideally measure ecosystem function rather than form; Bauder et al.’s (2009) HGM framework is an essential building block for this development in vernal pool restoration performance standards. In addition, to facilitate conformity of performance standard measurements, specific and repeatable methods for measuring metrics described in standards should be provided.
Lastly, we and others (Barbour et al. 2007; Solano HCP 2012; Calhoun et al. 2014) recommend a 5-year minimum for the length of monitoring, but long-term data collection is preferred as the system can change trajectory over time, as demonstrated by our study (Collinge & Ray 2009; Collinge et al. 2013).
Although we have framed the above principles largely in terms of vernal pool systems, they can be applied more broadly to restoration efforts in other ecosystems with varying restoration goals. We believe that poor performance standard guidelines are in part causing the current gap in the adaptive management cycle. By clarifying performance standards and related metrics and measurement methods, practitioners can more effectively use meaningful empirical project evaluation to inform next steps in management and improve long-term performance of restoration projects.
LITERATURE CITED
Barbour MG, Solomeshch AI, Buck JJ, Holland RF, Witham CW, MacDonald RL, Starr SL, Lazar KA (2007) Classification, ecological characterization, and presence of listed plant taxa of vernal pool associations in California.
Final Report, United States Fish and Wildlife Service Agreement/Study, Number 814205G238. University of California, Davis, California. http://w.vernalpools.net/documents/Barbour_et_al_2007.pdf (accessed 25 Sep 2014)
Barbour MG, Solomeshch AI, Holland RF, Witham CW, Macdonald RL, Cilliers SS, Molina JA, Buck JJ, Hillman JM (2005) Vernal pool vegetation of California: communities of long-inundated deep habitats. Phytocoenologia 35:177–200
Bauder ET, Bohonak AJ, Hecht B, Simovich MA, Shaw D, Jenkins DJ, Rains M (2009) A draft regional guidebook for applying the hydrogeomorphic approach to assessing wetland functions of vernal pool depressional wetlands in Southern California. San Diego State University, San Diego, California. http://www.bio.sdsu.edu/pub/andy/vernal pools/Bauder2009_07_App.pdf (accessed 10 Oct 2014)
Bernhardt ES, Sudduth EB, Palmer MA, Allan JD, Meyer JL, Alexander G, et al. (2007) Restoring rivers one reach at a time: results from a survey of U.S. river restoration practitioners. Restoration Ecology 15:482–493 Black C, Zedler PH (1998) An overview of 15 years of vernal pool restoration and construction activities in San Diego County, California. Pages195–205.
In: Witham CW, Bauder ET, Belk D, Ferren WR Jr., Ornduff R (eds) Ecology, conservation, and management of vernal pool ecosystems. Proceedings from a 1996 Conference. California Native Plant Society, Sacramento, California Bronner CE, Bartlett AM, Whiteway SL, Lambert DC, Bennett SJ, Rabideau AJ (2013) An assessment of U.S. stream compensatory mitigation policy:
necessary changes to protect ecosystem functions and services. Journal of the American Water Resources Association 49:449–462
Brooks RT, Hayashi M (2002) Depth-area-volume and hydroperiod relationships of ephemeral (vernal) forest pools in southern New England. Wetlands 22:247–255
Calhoun AJK, Arrigoni J, Brooks RP, Hunter ML, Richter SC (2014) Creating successful vernal pools: a literature review and advice for practitioners.
Wetlands 34:1027–1038
City of San Diego (2015) City of San Diego vernal pool habitat conservation plan. San Diego, California. http://www.sandiego.gov/planning/ programs/mscp/pdf/vphcp_prelimin_draft_march_2015.pdf (accessed 1 May 2015)
Cole CA (2002) The assessment of herbaceous plant cover in wetlands as an indicator of function. Ecological Indicators 2:287–293
Collinge SK (2003) Constructed vernal pool seeding experiment aids in recovery of Contra Costa goldfields (California). Ecological Restoration 21:316–317
Collinge SK, Ray C (2009) Transient patterns in the assembly of vernal pool plant communities. Ecology 90:3313–3323
Collinge SK, Ray C, Marty JT (2013) A long-term comparison of hydrology and plant community composition in constructed versus naturally occurring vernal pools. Restoration Ecology 21:704–712
CWMW (California Wetland Monitoring Workgroup) (2010) Tenets of a state wetland and riparian monitoring program. Sacramento, California. http://www.mywaterquality.ca.gov/monitoring_council/wetland_ workgroup/docs/2010/tenetsprogram.pdf (accessed 13 Oct 2014) De Weese JM (1998) Vernal pool construction monitoring methods and habitat replacement evaluation. Pages 217–223. In: Witham CW, Bauder ET, Belk D, Ferren WR Jr., Ornduff R (eds) Ecology, conservation, and management of vernal pool ecosystems. Proceedings from a 1996 Conference.
California Native Plant Society, Sacramento, California Doyle MW, Shields FD (2012) Compensatory mitigation for streams under the clean water act: reassessing science and redirecting policy. Journal of the American Water Resources Association 48:494–509
Gerhardt F, Collinge SK (2003) Exotic plant invasions of vernal pools in the Central Valley of California, U.S.A. Journal of Biogeography 30:
1043–1052
Gerhardt F, Collinge SK (2007) Abiotic constraints eclipse biotic resistance in determining invasibility along experimental vernal pool gradients. Ecological Applications 17:922–933 Harman W, Starr R, Carter M, Tweedy K, Clemmons M, Suggs K, Miller C (2012) A function-based framework for stream assessment and restoration projects. EPA 843-K-12-006. United States Environmental Protection Agency, Office of Wetlands, Oceans, and Watersheds, Washington, D.C.
Hobbs RJ, Harris JA (2001) Restoration ecology: repairing the earth’s ecosystems in the new millennium. Restoration Ecology 9:239–246
Keeley JE, Zedler PH (1998) Characterization and global distribution of vernal pools. Pages 1–14. In: Witham CW, Bauder ET, Belk D, Ferren WR Jr., Ornduff R (eds) Ecology, conservation, and management of vernal pool ecosystems. Proceedings from a 1996 Conference. California Native Plant Society, Sacramento, California
Palmer MA, Bernhardt ES, Allan JD, Lake PS, Alexander G, Brooks S, et al. (2005) Standards for ecologically successful river restoration. Journal of Applied Ecology 42:208–217
Palmer MA, Hondula KL (2014) Restoration as mitigation: analysis of stream mitigation for coal mining impacts in southern Appalachia. Environmental Science and Technology 48:10552–10560
Palmer MA, Hondula KL, Kock BJ (2014) Ecological restoration of streams and river: shifting strategies and shifting goals. Annual Review of Ecology, Evolution, and Systematics 45:247–269
Philippi TE, Simovich MA, Bauder ET, Moorad JA (2001) Habitat ephemerality and hatching fractions of a diapausing anostracan (Crustacea: Branchiopoda). Israel Journal of Zoology 47: 387–396.
Scheffer M (2009) Critical transitions in nature and society. Princeton University Press, Princeton, New Jersey
Solano Water County Agency (2012) Solano habitat conservation plan.
Solano County, California. http://scwa2.com/water-supply/habitat/solanomultispecies-habitat-conservation-plan (accessed 25 Sep 2014) Suding KN (2011) Toward an era of restoration in ecology: successes, failures, and opportunities ahead. Annual Review of Ecology, Evolution, and Systematics 42:465–487 USACOE (United States Army Corps of Engineers), USEPA (United States Environmental Protection Agency) (2008) Compensatory mitigation for losses of aquatic resources. Federal Register 73:19594–19705 March 2016 Restoration Ecology 151
Performance standards
USFWS (U.S. Fish and Wildlife Service) Region 1(2005) Recovery plan for vernal pool ecosystems of California and Southern Oregon.
Portland, Oregon. http://www.fws.gov/sacramento/es/Recovery-Planning/ Vernal-Pool/es_recovery_vernal-pool-recovery.htm (accessed 5 Oct 2014)
Williams BK (2011) Adaptive management of natural resources— framework and issues. Journal of Environmental Management 92:
1346–1353
Woolsey S, Capelli F, Gonser T, Hoehn E, Hostmann M, Junker B, et al. (2007) A strategy to assess river restoration success. Freshwater Biology 52:752–769
Wortley L, Hero JM, Howes M (2013) Evaluating ecological restoration success:
a review of the literature. Restoration Ecology 21:537–543 Zedler PH (1987) The ecology of southern California vernal pools: a community profile. U.S. Fish and Wildlife Service, Biological Report 85 (7.11), Pages 136
Coordinating Editor: Margaret Palmer Received: 9 June, 2015; First decision: 27 July, 2015; Revised: 30 November, 2015; Accepted: 30 November, 2015; First published online: 22 January, 2016 152 Restoration Ecology March 2016
The US Clean Water Act and Habitat Replacement:
Evaluation of Mitigation Sites in Orange County,
California, USA
MARK F. SUDOL
U.S. Army Corps of Engineers
Los Angeles District, Regulatory Branch
911 Wilshire Blvd., Suite 1100
Los Angeles, California 90017-3401, USA
RICHARD F. AMBROSE*
Environmental Science and Engineering Program
University of California, Los Angeles
Box 951772
Los Angeles, California 90095-1772, USA
ABSTRACT / Both permit requirements and ecological assessments have been used to evaluate mitigation success.
This analysis combines these two approaches to evaluate mitigation required under Section 404 of the United States Clean Water Act (CWA) and Section 10 of the Rivers and Harbors Act, which allow developers to provide compensatory mitigation for unavoidable impacts to wetlands. This study reviewed permit files and conducted field assessments of mitigation sites to evaluate the effectiveness of mitigation required by the US Army Corps of Engineers for all permits issued in Orange County, California from 1979 through 1993. The 535 permit actions approved during this period allowed 157 ha of impacts. Mitigation was required on 70 of these actions, with 152 ha of enhanced, restored, and created habitat required for 136 ha of impacts. In 15 permit actions, no mitigation project was constructed, but in only two cases was the originally permitted project built; the two cases resulted in an unmitigated loss of 1.6 ha. Of the remaining 55 sites, 55% were successful at meeting the permit conditions while 11% failed to do so. Based on a qualitative assessment of habitat quality, only 16% of the sites could be considered successful and 26% were considered failures. Thus, of the 126 ha of habitat lost due to the 55 projects, only 26 ha of mitigation was considered successful. The low success rate was not due to poor enforcement, although nearly half of the projects did not comply with all permit conditions. Mitigation success could best be improved by requiring mitigation plans to have performance standards based on habitat functions.
As a result of extensive wetland losses, there has
been increased attention paid to wetland management
and restoration worldwide. The United States has a
long history of wetland management regulations and
policies, principally based on the concept of mitigation.
Mitigation policy, established as a way to allow necessary development projects while minimizing environmental
impacts, was formalized in the National Environmental
Policy Act (NEPA) of 1969 and subsequent agency policies and memoranda of agreement (Berry and Dennison 1993). NEPA recognizes several different types of mitigation, but compensatory mitigation, which allows
resource losses to be compensated by the provision of
replacement resources, has drawn the most attention;
hereafter, “mitigation” is used to mean compensatory
mitigation.
Section 404 of the US Clean Water Act (CWA) includes mitigation as part of its regulation of impacts to wetlands (Kruczynski 1990, Ambrose 2000). Creation
and restoration of wetland habitat as mitigation has
received little support in the scientific literature (Race 1985, Zentner 1988, Kentula and others 1992, Holland
and Kentula 1992, DeWeese and Gould 1994, Miller
1995, Zedler 1996, Gilman 1998, Breaux and Serefiddin
1999, Brown and Veneman 2001). Recent criticisms of
mitigation projects include a lack of information on
wetland functions (Mitsch and Wilson 1996), insufficient understanding of the complexity of ecosystems, neglect of landscape-level interactions (Zedler 1996,
Kelly 2001), and inadequate compliance efforts by regulatory agencies. In spite of these problems, the use of mitigation continues because it remains part of the
CWA regulations. Many of these deficiencies are beginning to be identified and studied in the science of restoration ecology (Mitsch and Wilson 1996), but final solutions remain elusive (Ambrose 2000).
The US Army Corps of Engineers (Corps) regulates
activities within wetlands in the United States under
Section 404 of the CWA (all “waters of the United
KEY WORDS: Clean Water Act Section 404; Compensatory mitigation; Ecological assessment; Habitat creation; Mitigation success; Restoration; Riparian habitat; Wetlands; Orange County, California
*Author to whom correspondence should be addressed.
DOI: 10.1007/s00267-002-2787-3
Environmental Management Vol. 30, No. 5, pp. 727–734 © 2002 Springer-Verlag New York Inc.
States”) as well as Section 10 of the Rivers and Harbors Act (navigable waterways). The Corps regulates the
discharge of dredged or fill material into wetlands and waters of the United States. Placement of fill and drainage of wetlands has the greatest potential for direct and permanent loss of habitat. Corps permits in 1994 allowed 6964 ha of impacts while requiring 15,385 ha of habitat creation or restoration as mitigation for projectrelated impacts. Similarly, in 1993, 4696 ha of habitat were impacted with 6154 ha of mitigation required
(Studt 1994). Both of these analyses imply that more
wetlands are being restored or created than are being
impacted as a result of Corps-issued permits, implying a good degree of protection for wetland resources. However, more detailed reviews of permit files for a number of states have shown that mitigation requirements generally do not completely compensate for the losses allowed through Section 404 permitting (Holland and
Kentula 1992, Kentula and others 1992, Sifneos and
others 1992a,b Kelly 2001). Moreover, official Corps
wetland numbers (Studt 1994) reflect only hectares of
impacts and required mitigation, not the functions or
values of those hectares. Corps statistics represent the best possible case because they do not consider whether the mitigation is less than successful or is not completed.
In one of the earliest studies evaluating the success
of mitigation projects, Race (1985) identified problems with all of the eleven sites she evaluated in San Francisco Bay, although Harvey and Josselyn (1986) argued that the problems were not as severe as Race indicated.
A review of 18 wetland sites associated with San Francisco Bay Conservation and Development Commission (BCDC) permits judged 33% of the projects to be
successful (Demgen, 1988). A summary of work completed before 1989 (Josselyn and others 1990) found mixed results, with only one survey (Zentner 1988)
indicating a success rate of at least 50%. More recently, DeWeese and Gould (1994) found 50% of the projects
evaluated achieving at least 75% compliance with stated permit conditions. Similar results were found in a recent study of mitigation sites in Southern California (Allen and Feddema 1996), which identified a success
rate of 67%, again for compliance with permit conditions. Several studies have suggested that increased enforcement of mitigation permits would improve compliance with permit conditions (Zedler and Weller 1990, Holland and Kentula 1992, Sifneos and others
1992a, DeWeese and Gould 1994).
In this study, the effectiveness of the Corps program
allowing mitigation as replacement for lost habitat is
evaluated by examining one region in detail. All permits issued by the Los Angeles District Corps of Engineers for work within wetlands and waters of the United States in Orange County, California, were assessed. The assessment combined the two approaches, evaluation of
permit requirements and evaluation of ecological conditions, commonly used to evaluate mitigation success.
All permit files were examined to determine the extent
of permitted impacts and planned mitigation, as in
other studies (Holland and Kentula 1992, Kentula and
others 1992, Sifneos and others 1992a,1992b, Breaux
and Serefiddin 1999, Kelly 2001). Next, field investigations determined: (1) If the project had been constructed, and (2) the success of the mitigation site based on (a) a determination of compliance with specific permit conditions and (b) a qualitative assessment of habitat constructed. Comparing permit compliance
with habitat quality recognizes the distinction between compliance success—how well the projects met their
permit requirements—and ecological success—how
well they replace the functions of natural wetlands
(Quammen 1986).
Materials and Methods
Database and File Survey
Official files for permits issued in Orange County by
the Regulatory Branch of the Army Corps of Engineers
were screened to determine hectares of impacts and
mitigation using methods similar to those of Holland
and Kentula (1992) and Sifneos and others (1992a,b).
Official permit files are kept in the Los Angeles District for every permit action processed by the Corps since
the early 1900s. Each permit file includes project information, all incoming and outgoing correspondence, and pertinent maps, charts or photographs. A summary
of the information in these files is maintained on an
automated database that enables project managers to
track past permit actions, conduct database searches
using key words or phrases, and provide data on permit
processing statistics as required by the Office of the
Chief of Engineers. Permits issued, denied, or withdrawn in Orange County were identified using the automated database. Our search identified 840 files
associated with regulatory actions in Orange County
from January 1978 (date of the oldest permit with
mitigation requirements) to May 1993 (date of the
initial database query). The official record for each
permit was examined individually in the district office, although records on projects issued before 1980 had
been archived and could not be found. Official records
were screened to confirm that the permit action occurred in the county, that mitigation was included as a special condition of the permit, and to obtain informa728 M. F. Sudol and R. F. Ambrose tion on hectares of impacts, hectares of mitigation (if applicable), and Corps regulatory authority (Section 10 or 404). Initial screening of the official records showed that 305 permits actions were erroneously identified as occurring in Orange County and actually occurred in
other counties. The remaining 535 actions were completed in Orange County between 1979 and 1993, and 70 of these permits had mitigation requirements written as special conditions. Most permits concerned riparian habitat that was jurisdictional “waters of the United States” under Corps regulations, although they
may not have been “three-parameter” jurisdictional
wetlands. In this paper, we consider riparian habitat as a wetland habitat in the broad sense, even though it
may not meet the Corps regulatory requirements for a
wetland.
The specific type of mitigation was also determined
from the official record and verified during site inspections. These categories were defined as in Lewis (1990), with an additional category (passive revegetation) added to remove possible ambiguities. The categories are: (1) restoration: return to a preexisting condition; (2) creation: conversion of a persistent nonwetland (or riparian) habitat into wetland (or riparian) habitat (two subdivisions are recognized: artificial (irrigation required) or self-sustaining); (3) enhancement: increase in one or more functions due to intentional activities (i.e., plantings); (4) passive revegetation: allow a disturbed area to naturally revegetate without plantings. In addition, if the site was planned to be constructed on upland habitat and maintained
through the application of artificial irrigation (sprinklers or drip systems), this was noted and confirmed in the field.
Site Visits
We located the mitigation sites (important because
Corps staff had never visited many of the sites), recorded general site-specific information, and estimated the success of the mitigation project using two sets of criteria (Table 1). The first set is based on compliance with the success criteria established by the permit. Success criteria are written as special conditions of the permit or included in the applicant’s proposed mitigation plan and could be readily evaluated in the field.
Typical success criteria included percent survival of
transplanted plants, percent vegetation cover (all species), and/or the average height of vegetation after a specific period. Generally, the success criteria were proposed by the permit applicant’s consultant and approved by the Corps on a case-by-case basis. Thus, success criteria were different for nearly every mitigation project. The specific conditions identified in each permit were evaluated in the field to rate each site.
The second set of criteria concerns habitat quality,
including the density of the vegetation, plant species
diversity, presence of invasive plant species, and average tree height as compared to “native” habitat (defined as an area that is subject to normal regional hydrologic
conditions that has not been directly impacted by human activities). This approach, “best professional judgment,” is used frequently by resource agencies to assess habitat quality in permit decisions. Such qualitative
assessments have been the subject of much debate and
criticism (Harvey and Josselyn 1987). However, our
main purpose is to provide a rough estimate of ecological condition, so that the evaluation of mitigation success was not based solely on permit conditions.
Field visits were conducted from January 1994 to
January 1995. If no mitigation was found, then the
original project site was visited to learn if the project was ever built. If no project was built, the mitigation was classified as “no project.” If the project was built, another attempt was made to locate the mitigation site. If the mitigation site could not be found, the owner or
consultant was contacted in an attempt to determine if
mitigation was attempted. If the consultant/applicant
could not be contacted, the site was classified as “no
attempt.”
Table 1. Criteria for determining success of mitigation sitesa Category Permit conditions Qualitative estimate
Success Meets or exceeds all of the special conditions
associated with the Section 404 permit.
Habitat has similar spatial and species diversity as
minimally disturbed habitat.
Partial success Meets at least one of the special conditions associated with the Section 404 permit.
There is wetland or riparian vegetation on site, not
an upland site.
Failure Meets none of the special conditions associated with the Section 404 permit.
1. Majority of vegetation consists of upland species.
2. Less than 10% cover, all vegetation.
No attempt Project completed, no mitigation found.
No project Project not initiated, no mitigation needed. a
Criteria were applied during field visits to each mitigation site.
Mitigation Evaluation in Southern California 729
Results
The Corps issued 535 permits for work within Orange County from 1979 to 1993. Total impacts from these permits were 157 ha with 152 ha of mitigation
required, indicating a net loss of wetland habitat in
Orange County during this period. Only 70 of the 535
permits issued had requirements for mitigation; these
70 permits accounted for 87% (136 ha) of the area
impacted. The remaining 465 permits accounted for 21
ha of impacts, the result of many small projects authorized by the Corps that did not require mitigation. Most of these 465 permits were issued under Section 10 for
work within the bays and harbors of Orange County
that was considered to have minimal impacts on the
aquatic environment.
The 70 permits requiring mitigation consisted of 65
permits associated with Section 404 and five permits
associated with Section 10. Project-related wetland impacts and mitigation associated with the Section 404 permits in the county were divided into seasonally
flooded, intermittent riverine (riparian habitats), and isolated lacustrine systems (terminology after Cowardin and others 1979). The impacts associated with Section
10 permits involved intertidal estuarine habitats. Mitigation associated with three of the five Section 10 permits involved monetary contributions to the Newport Back Bay Ecological Preserve; only two Section 10 permits involved direct habitat restoration of estuarine habitat.
Field evaluations verified that 15 mitigation projects
had not been constructed (Table 2). These 15 sites
illustrate inaccuracies in estimates gleaned solely from permit files. Thirteen sites were classified as no project and two sites were classified as no attempt. The thirteen no-project sites would have had 8.1 ha of impacts and
13 ha of mitigation. Removal of these sites from the
acreage totals resulted in a reduction in total impacts from 136 to 128 ha and reduction in mitigation from
152 to 139 ha. Projects associated with the two noattempt sites were built, 1.6 ha of habitat was lost with no mitigation being constructed.
Evaluation of the remaining 55 mitigations found
that the majority of the sites were either successful or partially successful (Table 2). Using the project-specific permit conditions, 55% of the mitigation sites were
considered successful; in terms of acreage, 52% were
successful. Nineteen sites fulfilled at least one of the special conditions and thus were judged partially successful (35% of the sites, 32% of the acreage). Six sites were determined to be failures (11% of the sites, 15%
of the acreage). Using the criteria in the permits, the 55 projects resulted in the loss of 126 ha of habitat with only 72 ha of mitigation considered successful.
Qualitative evaluation of the mitigation sites considered fewer successful (16% or nine sites); in terms of acreage, 19% were successful. Thirty-two sites were considered partially successful (58% of the sites, 57% of the acreage). Fourteen sites were considered failures (26%
of the sites, 25% of the acreage). The 55 projects resulted in the loss of 126 ha of habitat with only 26 ha of mitigation considered successful. Thus, the evaluation
of success based on habitat yielded a lower success rate than the evaluation based on permit conditions.
Creation of habitat accounted for over two thirds of
the wetland mitigation acreage (87 ha of planned mitigation) associated with projects in Orange County (Table 3). The success of these mitigation projects as evaluated by the permit compliance showed that 80% (44 of the 55; 88% by acreage) were either successful or
partially successful. However, the qualitative evaluation indicates that approximately 20% (11 sites, 11 ha) of
these wetland mitigation projects were classified as failures. Except for passive revegetation projects, mitigation projects were judged 8%–13% less successful by the qualitative habitat evaluation compared to permit
compliance. Much of the success of the creation
Table 2. Success of mitigation projects based on permit compliance and habitat conditionsa Category
Permit compliance Habitat conditions
N
Impacts
(ha)
Mitigation
(ha) N
Impacts
(ha)
Mitigation
(ha)
Total permits 70 136 152 70 136 152
No project 13 8 13 13 8 13
No attempt 2 1.6 1.6 2 1.6 1.6
Projects evaluated 55 126 138 55 126 138
Success 30 91 72 9 42 26
Partial 19 24 45 32 63 78
Failures 6 11 21 14 21 34
a
Success criteria given in Table 1. Success was evaluated for the 55 mitigation projects actually initiated out of the 70 permits requiring mitigation. 730 M. F. Sudol and R. F. Ambrose
projects was due to irrigation. When creation sites requiring irrigation to maintain the habitat are removed from the analysis, the success rate declines to 14.5%
success (8 projects, 32% by acreage). There were no
restoration, enhancement, or passive revegetation sites requiring artificial irrigation.
Discussion
The evaluation of mitigation is often based on raw
acreage of wetland habitat created, restored, or enhanced as mitigation for projects negatively impacting aquatic or riverine habitat. For example, a survey of
wetland loss and mitigation by Tolman (1996), which
included official information from the US Fish and
Wildlife Service, National Resources Conservation Service, and the US Department of Agriculture, concluded that the US is gaining wetlands through construction of new habitat as mitigation. The Corps has published
statistics showing more area of mitigation being required than impacts allowed, again indicating a net gain of wetlands (US Army Corps of Engineers 1995).
This apparent gain in wetland acreage has been used to
support the policy of accepting mitigation for unavoidable impacts. However, these statistics do not consider the quality of the habitat created, restored, or enhanced as mitigation (Wilen 1995, Heimlich and Melanson 1995) or even whether the proposed mitigation is actually completed.
This study examined all permits issued within one
county in Southern California over a 15-year period to
characterize the success of past mitigation activities and assess the validity of statistics based on evaluation of permit requirements. Field investigations of Orange
County mitigation sites found that there was a net loss of wetland habitat, in spite of the fact that many mitigation sites did meet the success criteria included in the Corps permits. There were 126 ha of impacts from 55
permits with only 72 ha of successful and 45 ha of
partially successful mitigation, as evaluated by the success criteria in each permit. Even fewer of these sites were judged successful based on a qualitative assessment of the habitat quality, further reducing the acreage of effective mitigation. Even these estimates overstate the long-term success of the mitigation projects. In most cases, once a site meets the permit criteria, further maintenance or monitoring is not required, potentially
leading to additional losses as these sites decline due to loss of irrigation or invasion by nonwetland or nonriparian species.
The field assessment methods used in this study were
simple yet effective for evaluating important aspects of mitigation projects. Determining compliance success
(i.e., the degree to which mitigation projects meet the conditions of their permits) requires no special methodology but provides important insights beyond the information gleaned from an office review of permit
files. In this study, we found substantially less mitigation being provided than suggested from the permit files
alone. The qualitative assessment of ecological success, while no replacement for a detailed ecological assessment, was simple to perform and demonstrated that many mitigation projects judged successful in terms of
permit conditions were still resulting in resource losses.
Routine application of simple field assessment methods, such as employed in this study (see also Ainslie 1994, Stein and Ambrose 1998, Yoder and Rankin
1998), would provide a more accurate assessment of the
success of Section 404 mitigation than simple permitderived summaries of acreage.
Previous studies have called for increased compliance efforts by the Corps as a means to improve mitigation success (Zedler and Weller 1990, Holland and Kentula 1992, Sifneos and others 1992a, DeWeese and
Gould 1994). Better enforcement could certainly have
prevented the losses associated with the two permits in which no mitigation was attempted, but this was a minor contribution (1.6 ha) to the overall wetland losses identified in this study. Enforcement might also have
improved the success of the 11% of the projects judged
Table 3. Summary of mitigation success by mitigation type Mitigation type
Total
(ha)
Creation
(ha)
Restoration
(ha)
Enhancement
(ha)
Passive
revegetation
(ha)
Permitted impacts 86 22 6.5 11 126
Planned mitigation 87 28 9.7 13 138
Successful or partially successful by permit conditions 77 20 9.3 11 117 Successful or partially successful by qualitative evaluation 68 17 8.5 11 104 Successful or partially successful by qualitative evaluation and without artificial irrigation
32 17 8.5 11 67
Mitigation Evaluation in Southern California 731
failures and the 35% judged partially successful. However, 54% of projects studied met all of their permit conditions. If a site meets the conditions for success in its permit, the Corps approves the mitigation project
and can take no further action. In Orange County, the
Corps could not seek changes to more than half of the
mitigation sites because the permit conditions were
met.
Artificial irrigation has become a point of contention within several Corps districts. Applicants frequently propose artificial irrigation of upland sites as mitigation for impacts to wetlands and riparian habitat.
In the Los Angeles District, these projects have historically been accepted because of the difficulty of finding suitable areas for mitigation. The Corps, in the analysis of less-damaging project alternatives, cannot require an applicant to construct mitigation that is considered
prohibitively expensive or not practical using current
technology. Consequently, the Corps cannot require an
applicant to purchase property for mitigation if the cost of that land makes the entire project significantly more expensive. As a result, mitigation is usually performed on areas already owned by the applicant (or on public
lands) that lack wetland hydrology and do not support
wetland vegetation. Applicants propose various plans to provide water to the site under the assumption that the addition of water will support the planting and establishment of wetland or riparian vegetation.
Twelve of the 22 habitat creation sites in Orange
County were upland areas that had been cleared and
graded, with trees and shrubs planted in rows to facilitate installation and maintenance of the artificial irrigation system. These sites obviously do not resemble or function like natural riparian habitats; rather, they were designed to achieve the performance criteria established in their permits, which include tree survival rates, tree height, and percent of canopy cover. Not
surprisingly, the permit conditions have driven the design of these sites. Permitees desire a mitigation project that will meet the Corps’ mitigation requirements at
reasonable cost, and a field of regularly spaced and
irrigated trees satisfies these needs. Unfortunately, without natural hydrologic processes, which are essential for natural recruitment and the maintenance of species
diversity (Warner and Hendrix 1985, Faber and others
1989), these sites will never be able to support a selfsustaining riparian community. In addition, continued artificial irrigation is not included as a specific permit condition in many cases, so the water can be turned off once the success criteria are met. In these cases, the
vegetation will almost certainly revert back to an upland community. Because 36 ha of the mitigation projects
were irrigated, it is clear that the policy of allowing irrigation of upland areas has, by itself, ensured that the goal of no net loss of wetland habitat was not met.
While this practice has been curtailed in some districts, the practice of relying on vegetation parameters as the primary criteria for success continues (Breaux and
Serefiddin 1999).
Another practice recently questioned by resource
agencies is the appropriateness of allowing enhancement of riverine wetland habitat (rather than creation or restoration) as mitigation. This type of mitigation
does not provide new replacement habitat. Riverine
habitat in Southern California is in a constant state of flux due to the influence of the annual winter rains and associated flooding (Warner and Hendrix 1985, Faber
and others 1989). Riverine zones are areas of natural
scour and recolonization by specific plant communities; enhancement of an area under the influence of natural processes may alter the timing of succession but does not provide a net gain of habitat. If the goal of the enhancement is to increase the suitability of an area for an endangered species, then it might be considered
suitable mitigation, but there were no such mitigation
plans in Orange County. Allowing the enhancement of
wetlands as mitigation results in a net loss of habitat area, although the mitigation area may be protected
from future development through the use of deed restrictions or conservation easements.
The results of this study support previous findings
(GAO 1988, Kentula and others 1992, Race and Fonseca 1996) that a net loss of wetlands in the United States is continuing under current regulations. If the
national success rate is similer to that of Orange County’s, then thousands of hectares of wetlands are lost annually because of inadequate mitigation. In this
study, enforcement could have prevented some mitigation failures, although more than half of the projects met all of their performance standards. More importantly, improved mitigation and monitoring plans would rectify problems with inadequate mitigation
plans and inappropriate mitigation success criteria.
More recent plans contain more clearly defined mitigation objectives. Vague goals and objectives hindered earlier evaluations of mitigation success (DeWeese and
Gould 1994), whereas the projects we reviewed had
permit conditions that could be used to assess compliance. However, our results demonstrate that meeting regulatory requirements is not sufficient to protect wetland resources. Because permit conditions generally did not focus on replacing the functions lost when
natural riparian habitats were impacted, even “successful” projects (i.e., those meeting all permit conditions) could result in resource losses. Without permit conditions that better reflect ecological functions, mitigation 732 M. F. Sudol and R. F. Ambrose
efforts will allow the continued loss of wetland habitat and the ecosystem services they provide.
Acknowledgments
The authors thank the US Army Corps of Engineers,
Los Angeles District, Regulatory Branch for support,
access to files, and the ability to visit the mitigation sites.
Dr. Richard Schubel and Mr. Mark Durham of the
Corps provided the foresight and desire to evaluate past mitigation practices, which led to changes in mitigation policy in the Los Angeles District. We also thank Mr.
Henry Watson for providing transportation services for
most of the site evaluations. Finally, we thank Ms. Karleen Boyle and Drs. L. Donald Duke, Antony Orme, Philip Rundel, Mary Kentula, and Joy Zedler for editorial review and technical evaluation.
Literature Cited
Ainslie, W. B. 1994. Rapid wetland functional assessment—its role and utility in the regulatory arena. Water, Air and Soil Pollution 77:433–444.
Allen, A. O., and J. J. Feddema. 1996. Wetland loss and substitution by Section 404 permit program in Southern
California, USA. Environmental Management 20:263–274.
Ambrose, R. F. 2000. Wetland mitigation in the United States:
assessing the success of mitigation policies. Wetlands (Australia) 19:1–27.
Berry, J. F., and M. S. Dennison. 1993. Wetland Mitigation.
Pages 278–303 in M. S. Dennison and J. F. Berry (eds.), Wetlands: Guide to science, law and technology. Noyes
Publications, Park Ridge, New Jersey.
Breaux, A., and F. Serefiddin. 1999. Validity of performance criteria and a tentative model for regulatory use in compensatory wetland mitigation permitting. Environmental Management 24(3):327–336.
Brown, S. C., and P. L. M. Veneman. 2001. Effectiveness of compensatory wetland mitigation in Massachusetts, USA.
Wetlands 21:508–518.
Cowardin, L. M., V. Carter, F. C. Golet, and E. T. Laroe. 1979.
Classification of wetlands and deepwater habitats of the United States. US Fish and Wildlife Service, Washington, DC, 103 pp.
Demgen, F. C. 1988. A review of eighteen mitigation sites in the San Francisco Bay Region. Pages 318–322 in J. Kusler, D. Daly, and G. Brooks (eds.), Proceeding of the National Symposium: Urban Wetlands. Association of Wetland Managers.
DeWeese, J., and C. Gould. 1994. An evaluation of selected wetland creation projects authorized through the Corps of Engineers Section 404 Program. US Department of the
Interior, Fish and Wildlife Service. Sacramento Field Office, Sacramento, California, 90 pp.
Faber, P. A., Keller, E., Sands, A., and B. M. Massey. 1989. The ecology of riparian habitats of the southern California coastal region: A community profile. US Fish and Wildlife Service Biological Report 85(7.27), 152 pp.
GAO (General Accounting Office). 1988. Wetlands: The
Corps of Engineers’ administration of the Section 404 Program. US General Accounting Office, Washington, DC, 122 pp.
Gilman, E. L. 1998. Nationwide permit program: unknown
adverse impacts on the commonwealth of the Northern
Mariana Islands’ wetlands. Coastal Management 26:253–277.
Harvey, T. H., and M. N. Josselyn. 1986. Wetland restoration and mitigation policies: Comment. Environmental Management 10:567–569.
Heimlich, R., and J. Melanson. 1995. Wetlands lost, wetlands gained. National Wetlands Newsletter 17:23–25.
Holland, C. C., and M. E. Kentula. 1992. Impacts of Section 404 Permits requiring compensatory mitigation on wetlands in California. Wetlands Ecology and Management 2:157–169.
Josselyn, M., J. B. Zedler, and T. Griswold. 1990. Wetland mitigation along the Pacific Coast of the United States.
Pages 3–36 in J. A. Kusler and M. E. Kentula (eds.), Wetland creation and restoration, the status of the science. Island Press, Washington, DC.
Kelly, N. M. 2001. Changes to the landscape patterns of coastal North Carolina wetlands under the Clean Water Act.
Landscape Ecology 16:3–16.
Kentula, M. E., J. C. Sifneos, J. W. Good, M. Rylko, and K.
Kunz. 1992. Trends and patterns in Section 404 permitting requiring compensatory mitigation in Oregon and Washington, USA. Environmental Management 16:109–119.
Kruczynski, W. L. 1990. Mitigation and the Section 404 program: A perspective. Pages 549–554 in J. A. Kusler and M. E.
Kentula (eds.), Wetland creation and restoration, the status of the science. Island Press, Washington, DC.
Lewis, R. R. 1990. Wetlands restoration/creation/enhancement terminology: Suggestions for standardization. Pages 417–422 in J. A. Kusler and M. E. Kentula (eds.), Wetland creation and restoration, the status of the science. Island Press, Washington, DC.
Miller, J. 1995. Man-made wetlands’ future starts to dry up. Los Angeles Times, East Ventura County Edition, 9 July 1995, p.
B1.
Mitsch, W. J., and R. F. Wilson. 1996. Improving the success of wetland creation and restoration with know-how, time, and self-design. Ecological Applications 6:77–83.
Quammen, M. L. 1986. Measuring the success of wetlands
mitigation. National Wetlands Newsletter 8(5):6–8.
Race, M. S. 1985. Critique of present wetland mitigation policies in the United States based on an analysis of past restoration projects in San Francisco Bay. Environmental Management 9:71–82.
Race, M. S., and M. S. Fonseca. 1996. Fixing compensatory mitigation: what will it take? Ecological Applications 6:94–101.
Sifneos, J. C., E. W. Cake, Jr., and M. E. Kentula. 1992a. Effects of Section 404 permitting on freshwater wetlands in Louisiana, Alabama, and Mississippi. Wetlands 12:28–38.
Sifneos, J. C., M. E. Kentula, and P. Price. 1992b. Impacts of Section 404 permits requiring compensatory mitigation of Mitigation Evaluation in Southern California 733
freshwater wetlands in Texas and Arkansas. The Texas Journal of Science 44:475–485.
Stein, E. D., and R. F. Ambrose. 1998. A rapid impact assessment method for use in a regulatory context. Wetlands 18:393–408.
Studt, J. F. 1994. Wetland impact acreage analysis. Department of the Army, US Army Corps of Engineers, Memorandum. 22 March.
Tolman, J. 1996. Achieving no net loss. National Wetlands Newsletter 17(3):5–7.
US Army Corps of Engineers. 1995. Section 404 of the Clean Water Act and Wetlands: Special Statistical Report. US
Army Corps of Engineers, Regulatory Branch, July.
Warner, R. E., and K. M. Hendrix. 1985. Riparian resources of the central valley and California desert. California Department of Fish and Game, Sacramento, California.
Wilen, B. O. 1995. The folly of the numbers game. National Wetlands Newsletter 17:8–10.
Yoder, C. O., and E. T. Rankin. 1998. The role of biological indicators in a state water quality management process.
Environmental Monitoring and Assessment 51:61–88.
Zedler, J. B. 1996. Coastal mitigation in Southern California:
the need for a regional regional restoration strategy. Ecological Applications 6:84–93.
Zentner, J. J. 1988. Wetland restoration success in coastal California. Page 216–219 in J. Zelazny and J. S. Feierabend (eds.), Increasing our wetland resource: proceedings of a conference.
National Wildlife Federation, Washington DC, 324 pp.
734 M. F. Sudol and R. F. Ambrose
February 21, 2023
Sent via email
Brent Cooper
City of American Canyon
4381 Broadway Street, Suite 201
American Canyon, CA 94503
[email protected]
Re: Giovannoni Logistics Center Project Final Environmental Impact Report-Volume 2 (SCH # 2021010104)
Dear Mr. Cooper:
These comments are submitted on behalf of the Center for Biological Diversity (the “Center”) regarding the Giovannoni Logistics Center. The Center has reviewed the Final Environmental Impact Report Volume 2 issued on February 7, 2023 (“Revised Final EIR” or “RFEIR”) closely and remains concerned about the inadequate analysis of the new road that bisects the site, as well as the still-inappropriate GHG threshold and errors in the GHG analysis, the inaccurate wetland analysis, and the conclusory dismissal of impacts to Swainson’s hawk habitat. The Center urges the City Council to delay approval of the Project until the EIR can be revised and recirculated.
The Center is a non-profit, public interest environmental organization dedicated to the protection of native species and their habitats through science, policy, and environmental law.
The Center has over 1.7 million members and online activists throughout California and the United States. The Center has worked for many years to protect imperiled plants and wildlife, open space, air and water quality, and overall quality of life for people in Napa County and the City of American Canyon.
I. THE NEW ROAD THROUGH THE PROJECT SITE—AND THROUGH THE WETLAND PRESERVE—REQUIRES RECIRCULATION OF THE EIR.
“Public participation is an essential part of the CEQA process.” (CEQA Guidelines § 15201.) Given the public’s vital role in the CEQA process, a lead agency must recirculate an EIR for additional of public review and comment when, after the close of the comment period, “new significant information” shows that the EIR’s analysis was fundamentally inadequate and conclusory in nature such that that meaningful public review and comment were precluded. (CEQA Guidelines § 15088.5(a)(4).) New information is significant where the EIR is changed in February 21, 2023
a way that “deprives the public of a meaningful opportunity to comment” on a substantial adverse effect or feasible (but unimplemented) mitigation measures or alternatives. (CEQA Guidelines § 15088.5(a) [emphasis added].) Where new information demonstrates the inadequacy of the EIR’s analysis, recirculation is required “so that the public is not denied an ‘opportunity to test, assess, and evaluate the data and make an informed judgment as to the validity of the conclusions to be drawn therefrom.’” (Save Our Peninsula Committee v. Monterey County Bd. of Supervisors (2001) 87 Cal.App.4th 99, 131 [emphasis added].) The City did not disclose the gravel road that bisects the site, including the City’s proposed mitigation. The City did not acknowledge the existence of the road until commenters discovered it. The gravel road – and its effects on the City’s mitigation plan – is significant new information. Under California law, this EIR must be recirculated.
A. The City’s Choice of Baseline Is Not Supported by Substantial Evidence Because It Ignores the New Road that Is Necessary for a Realistic Baseline.
An EIR must provide an accurate and complete description of the “baseline” existing environmental conditions against which a project’s impacts are evaluated. (Neighbors for Smart Rail v. Exposition Metro Line Construction Authority (2013) 57 Cal.4th 439, 447-48; Guidelines § 15125.) “The public and decision makers are entitled to the most accurate information on project impacts practically possible, and the choice of a baseline must reflect that goal.” (Id. at 455.)
The City is claiming that the road that existed on the site when the EIR was finalized is not a necessary part of the environmental baseline because the road was not part of the physical condition of the site on January 12, 2021, the date the NOP was issued. (RFEIR 2-11.) However, there is no “uniform, inflexible rule” that the most accurate baseline will always be the condition as of the date of the NOP. (Neighbors for Smart Rail, 57 Cal.4th at 449.) Instead, an agency must decide “how the existing physical conditions without the project can most realistically be measured.” (Communities for a Better Environment v. South Coast Air Quality Management Dist. (2010) 48 Cal.4th 310, 328.) The agency’s decision must be supported by substantial evidence showing that the chosen baseline represents “the most accurate picture practically possible of the project's likely impacts.” (Ibid.)
By choosing to ignore a road that was already slated to be built when the NOP was issued and was built soon after, the City chose a baseline that was not the most accurate picture practically possible. It was already clear that the road would be built. As the City has pointed out, the plan to build the road was conclusively documented in the environmental review documents for the Napa Logistics Park Project and various permits. (RFEIR at 2-10). Therefore, any attempt to choose an accurate baseline would necessarily include the road that was understood to be imminent. Instead, the City chose a baseline that did not reflect the goal of giving the public and decision makers the most accurate information, even though that information was clear and available to the City at the time.
February 21, 2023
B. Because the EIR Did Not Disclose or Analyze the New Road, its Analysis of the Wetland Impacts and Mitigation is Not Supported by Substantial Evidence.
The City did not consider the road’s existence in its analysis of the wetland mitigation. The wetland preserve is the Project’s only mitigation for impacts to approximately 163 acres of habitat. Because the new road cuts directly through the land proposed as a wetland preserve, the road makes all the discussion of the wetland preserve, including maps of the preserve, inaccurate. (See, e.g. DEIR at Exhibit 3.3-3.) The City has now, for the first time, admitted that the wetland preserve will contain this development, which brings with it vehicle trips, emissions, and human interference. However, it argues that the road will not impact the wetland mitigation scheme. (RFEIR at 2-12-15)
New information cannot remedy a DEIR’s inadequate analysis of a project’s impacts in the first instance. (Ukiah Citizens for Safety First v. City of Ukiah (2016) 248 Cal.App.4th 256, 266-
67.). By disclosing the road and arguing that it is not significant only now, in the second FEIR, the City has given the public no chance to review or respond to significant new claims regarding what has always been a central concern in this environmental review process—whether or not the 45-acre preserve can adequately mitigate for 163 lost acres of habitat. The connectivity of the preserve and the degree to which it is harmed by the edge effects of the development that hems it in are crucial factors in its effectiveness and they require a fuller analysis.
II. THE WETLAND ANALYSIS IS INADEQUATE AND IMPROPERLY
DEFERRED.
A. The EIR Proposes Mitigation Ratios that Are Too Low and Not Substantially Likely to Be Successful.
Wetland mitigation can be addressed in a number of ways. Here, the proposed project proposes to mitigate through an “in-kind,” which is to protect existing wetlands outside of the proposed project impact. Typically, as stated in our prior comments, adequate mitigation is provided by the project proponent in a way to assure future protections through a much higher mitigation ration than 1:1. The State in compliance with the requisite Lake and Streambed Alteration permit typically requires a 3:1 ratio of preserved wetlands to impacted wetlands. As stated in our original comments, a 3:1 mitigation ratio is typical - three acres of preservation of existing, ecologically functional wetlands for every acre of wetlands impacted by the proposed project. Unfortunately, even this strategy has not achieved the no net loss of wetlands. (Sudol and Ambrose 2002; Ambrose 2000; Ambrose, Callaway, and Lee 2006; Calhoun et al. 2014; Schlatter, Faist, and Collinge 2016). If the proposed project does not choose to mitigate for the wetlands impacts “in-kind”, it then proposes to mitigate through enhancement of degraded wetlands to a point where self-sustaining ecological function is achieved, or to create wetlands.
These types of mitigation are fraught with even less successful outcomes. (Ibid.) Both of these options require higher levels of mitigation because they are so often unsuccessful in reaching success criteria of a fully self-sustaining wetlands over time i.e. unsuccessful in reaching a nonet-loss-of-wetlands standard. The approaches of enhancement or creation are often time February 21, 2023
consuming (along with a significant monetary investment), causing a temporal impact where the development impact eliminates wetland habitat, yet the enhancement or creation wetlands have not matured to be fully functional and the displaced plants and animals reliant upon the impacted wetland remain impacted due to habitat loss.
B. The Mitigation Measures for Wetlands Are Improperly Deferred.
The RFEIR admits that the earth disturbance associated with the new road impacts areas proposed for wetland creation and that the utility lines under the road now make the road area unsuitable for wetland creation. (RFEIR at 2-14-15) However, the City does not expect this to impact the preserve because the location of the created wetlands will be shifted in an unspecified way. (RFEIR at 2-14.) This improperly deferred mitigation violates CEQA. (see San Joaquin Raptor Rescue Center v. County of Merced (2007) 149 Cal.App.4th 645, 670 [EIR inadequate where the success or failure of mitigation efforts “may largely depend upon management plans that have not yet been formulated, and have not been subject to analysis and review within the EIR”]). In the limited circumstances in which deferred mitigation is appropriate, the agency must meet all of the following elements: (1) practical considerations prevented the formulation of mitigation measures during the planning process; (2) the agency committed itself to developing mitigation measures in the future; (3) the agency adopted specific performance criteria prior to project approval; and (4) the EIR lists the mitigation measures to be considered, analyzed, and possibly incorporated into the mitigation plan. (See POET, LLC v. State Air Resources Bd. (2013) 218 Cal.App.4th 681, 736-37 [review denied].) Here, the EIR fails to meet these criteria. There is also no reason the location of the created wetlands cannot be provided to the public now. In fact, the DEIR did attempt to map out the created wetlands. (DEIR at Exhibit 3.3-4.) The City is only claiming that the created wetlands will be “fieldfit,” and therefore maps of where they will be created are unnecessary, after commenters discovered that the provided maps could not be accurate. (RFEIR 2-14.) The City’s first failure to include accurate maps makes it more important that those details be disclosed to the public now, not less.
III.THE PROJECT’S GHG EMISSIONS STILL HAVE NOT BEEN CORRECTLY ANALYZED.
A. The RFEIR Does Not Support the Conclusions of the Fehr & Peers Study with Substantial Evidence.
The City continues to rely on the Fehr & Peers study to support the claim that the Project VMT from trucks will be 15% below the regional average, despite the serious deficiencies in that study raised in earlier comments. In response to these concerns, the RFEIR says that the study “used locally collected data” and “is project-specific.” (RFEIR 2-27.) In fact, the trip generation rates were collected in San Joaquin County and are used for this project with no justification for the assumption that an average from San Joaquin County is an accurate proxy for this Project.
Overall, the analysis is the opposite of project-specific. The study uses average trip length from other warehouses in Napa County without justifying the assumption that the Project will have similar tenants or similar use intensity, combined with the average trip generation from San Joaquin County. Because there is nothing project-specific about this analysis, it would conclude that every future warehouse in Napa County will, almost by definition, have less-than-significant VMT impacts. An analysis that has this effect cannot be supported by substantial evidence.
February 21, 2023
C. The RFEIR Does Not Support the Choice of GHG Significance Thresholds with Substantial Evidence.
The RFEIR retains the same inadequate GHG threshold that inappropriately disregards the significant truck trips that will be generated by the Project. The City says that the significance thresholds are based on BAAQMD’s 2022 thresholds, which are designed to be in line with the “fair share” concept and the State’s 2030 emissions targets, as opposed to the 2017 thresholds, which would be inappropriate because they are in line with earlier emissions thresholds. (RFEIR 3-272.) The City is thus making the confusing claim that although the Project would be significant under thresholds that were based on the goal of returning to 1990 levels of GHG emissions, the Project is not significant under thresholds based on the more ambitious goal of reducing GHGs to 40% below the 1990 levels. The reason for this inconsistency is that the 2022 thresholds are for office buildings and are not meant for industrial projects, and the 2017 thresholds, under which the impacts are significant, were the last thresholds BAAQMD did create for industrial uses. The fact that the 2017 thresholds are based on older emissions goals might be a rationale for using a higher threshold now, but it cannot be a rationale for using a lower one.
D. The RFEIR’s Explanation of Employee Commute Trip Generation is Still Contradictory and Not Supported by Evidence.
As pointed out in the Center’s earlier letter, the EIR’s estimates of how many trips will be generated by employee commutes are much too low to be consistent with the estimated employment. The RFEIR addresses this by adding new claimed employment numbers that are themselves not supported by substantial evidence, and further, still are not consistent with the trip generation estimate. It revises the project’s estimated employment to say that only 2,400 people will be employed on site, and the rest of the claimed 3,643 total employees come from the supposition that that the project “may well” indirectly create economic growth, and thus, create 1,243 jobs additional to the jobs created at the warehouse. (RFEIR at 3-274.) Those jobs will be indirectly associated with commerce around the project site and the employees will work off-site, which is the explanation given for why their commutes are not counted in the trip generation numbers for the Project. The RFEIR says this distinction between 2,400 on-site employees and 1,243 additional “indirect” employees is “as described in the Draft EIR at page 2-8.” In fact, the Draft EIR at page 2-8 only says that “[t]he proposed project would employ…an estimated 3,643 workers at buildout.” It does not say that those 3,643 employees will be “associated with commerce, both directly and indirectly, around the project site,” or that any of them will be employed off-site. Those claims are made for the first time in the revised FEIR.
The claims are also made without support. The response offers nothing to explain why it estimates 2,400 on-site workers, and the estimate of 1,243 additional workers is even more unclear because the City has not said anything about how it projected the Project’s indirect impacts on economic growth in the area, much less how many jobs that projected growth will generate. Further, because the EIR must analyze all direct and indirect impacts of the project, if 1,243 jobs will be indirectly generated by the project, the EIR must analyze whether those employees will commute to wherever their newly generated jobs will be, and, if so, the impacts February 21, 2023
of those commutes. By counting the 1,243 jobs as part of the economic benefit of the project because they “may well” be generated by economic growth that the Project “indirectly” causes, but assuming they will have no VMT/GHG impacts, the EIR continues to mislead the public with conflicting claims.
Moreover, the EIR’s claim that 2,400 on-site employees will generate only 2,832 trips per day is not supported by substantial evidence. A reasonable set of assumptions would be that each on-site employee will drive to and from work five days a week. Under those assumptions, estimated trip generation would be 3,429 trips per day. Somehow the RFEIR estimates almost 600 fewer trips per day. Whatever assumptions the RFEIR is using to conclude that on-site employees will drive to work substantially less than five times a week are not explained to the public and the conclusion is supported by no evidence.
IV.SWAINSON’S HAWK ARE A SPECIAL STATUS SPECIES, THE IMPACTS TO WHICH ARE PRESUMED TO BE SIGNIFICANT.
In addition to being protected under CEQA, many bird species affected by industrial development are protected under various other federal and state laws, including the federal and state Endangered Species Acts, federal Migratory Bird Treaty Act, federal Bald and Golden Eagle Protection Act, and multiple provisions of the California Fish and Game Code. As relevant to this Project, for example, the site is foraging and nesting habitat for the Swainson’s hawks, which is listed as threatened species under the California Endangered Species Act (CESA), Fish and Game Code section 2050 et seq., and thereby protected from “take” under this state law. (Fish & G. Code, § 2080.)1
The CEQA Guidelines indicate that a Project can be expected to have significant impacts to biological resources if the Project has a substantial adverse effect, either directly or through habitat modifications, on any species identified as a candidate, sensitive, or special status species in local or regional plans, policies, or regulations, or by the California Department of Fish and Game or U.S. Fish and Wildlife Service. (CEQA Guidelines, Appendix G, subd. IV(a).) Accordingly, the DEIR itself indicates that the Project’s impacts will be significant if it will “have a substantial adverse effect, either directly or through habitat modifications, on any species identified as a candidate . . . species . . . by the California Department of Fish and Game.” (DEIR at 4.3-18; see also CEQA Guidelines § 15065(a)(1) [when performing an initial study, agencies shall make a mandatory finding of significance where a proposed project has the potential to substantially reduce the number or restrict the range of a listed species].) Consequently, the Project’s impacts to Swainson’s hawk must be considered significant and fully 1 Fish and Game Code section 86 defines “take” as to “hunt, pursue, catch, capture, or kill,” or to attempt to do any of these things. Fish and Game Code sections 3503.5 and 3513 further prohibit take of any raptors and their nests and eggs, and prohibit take of all birds protected under the federal Migratory Bird Treaty Act, respectively. (See also id., § 3800, subd. (a) (prohibiting take of all non-game birds, except as provided in the Fish and Game Code).) February 21, 2023
evaluated. The City has failed to do, disregarding the evidence, studies, and recommendations of the California Department of Fish and Wildlife (CDFW).
The loss of foraging habitat is the primary threat to Swainson’s hawk, and the City does not propose any mitigation for the significant impacts to Swainson’s hawk foraging habitat, which are presumed to be significant. The City proposes to develop approximately 154.7 acres of valuable Swainson’s hawk habitat into 2.4 million square feet of industrial/warehouse space and associated improvements. (ES at 1-1.) The Project’s proposed warehouse will require the removal of vegetation from the site prior to the start of construction (DEIR at 3.3-38.), which will necessarily include any foraging or nesting habitat located in the Project footprint.
Swainson’s hawks were observed foraging on the site, and nests have been observed near the site. (RFEIR at 2-2.) The City admits that the Project will result in a loss of suitable habitat. (Ibid.) Nevertheless, the City still refuses to acknowledge any potentially significant direct or indirect impacts associated with the destruction or adverse modification of the Swainson’s hawk habitat.
The City makes three claims, all flawed, for why it need not mitigate habitat impacts, in blatant disregard of CEQA, the California Endangered Species Act (CESA), and CDFW guidance.
First, the City disregard CDFW’s comments on the EIR and the findings in CDFW’s 2016 5-Year Review, assuming that the loss of foraging habitat is not significant because Swainson’s hawk populations are “rebounding.” Under CEQA, a species is presumed to be endangered, rare, or threatened when it is listed under California endangered species regulations (14 Cal Code Regs §§670.2, 670.5; 14 Cal Code Regs §15380.) Swainson’s hawk is listed under CESA. The City cannot supplant its own cherry-picked science of rebounding numbers to conclude the Swainson’s hawks are no longer subject to protection under CEQA or CESA. To the contrary, both laws require that the City mitigate the loss of habitat. No less, the very study upon which the City relies cautions the reader that urban development and loss of grasslands (both which will occur as a result of this Project) will continue to threaten the Swainson’s hawk. (RFEIR at 2-6.)
Second, the City seems to imply that the existence of other suitable habitat in the vicinity of the Project frees the City of its obligation to mitigate the loss of habitat on site. (RFEIR at 2-
2.) Swainson’s hawks are known to have high habitat site fidelity, meaning they return to the same site year after year. (Estep 1989, Woodbridge et al. 1995). The existence of other habitat does not absolve the Project of the requirement to mitigate the impacts of this project, as CEQA requires. Furthermore, the City fails to consider the several other large development projects near the proposed Project which also have not been required to mitigation for Swainson’s hawk foraging habitat. (CDFW December 2, 2022 letter.) As CDFW notes, the Project poses a clear cumulative impact to Swainson’s hawk in the area, and the existence of remaining habitat does not eliminate that impact.
Third, the City concludes that the appropriate baseline for assessing Swainson’s hawk foraging habitat should consider the increased Swainson’s hawk populations. Assuming this is true, it cuts against the City’s ultimate conclusion that the loss of habitat does not pose February 21, 2023
significant impacts. If this area hosts a rebounding population, then the site’s value for supporting Swainson’s hawk foraging habitat increases.
Finally, the City assumes – without evidence – that a 45-acre wetland preserve on-site would serve as suitable Swainson’s hawk habitat on-site, thereby mitigating some of the lost habitat. (RFEIR at 2-2, 2-3.) The City has no evidence in the record demonstrating that the manufactured wetland preserve – which as discussed above, includes an undisclosed easement road – will mitigate the impacts to Swainson’s hawk habitat. No less, the City cannot avoid studying impacts to Swainson’s hawk in the first instance by proposing a plan to mitigate presumed impacts, unless the mitigation measures and mitigation performance standards are identified. San Joaquin Raptor Rescue Ctr. v County of Merced (2007) 149 Cal.App.4th 645, 671. The EIR’s wetlands mitigation contains no such performance standards for Swainson’s hawk.
In sum, we recognize that Swainson’s hawks have recovered somewhat in central and northern California due to increased protections for the species because of its listing as threatened. Swainson’s hawks’ State-threatened status still requires adequate mitigation for the species, specifically mitigation for impacts to its foraging habitat. The FEIR remains inadequate in this respect because it fails to mitigate for the 163 acres of foraging habitat and instead relies on the 45 acres of the “wetland preserve.” At a minimum, the 45 acres of the proposed “wetland preserve” needs to be evaluated for adequacy of providing foraging habitat and additional acreage needs to be secured in perpetuity as foraging habitat for Swainson’s hawks, as has been done for other developments in Swainson’s hawk foraging habitat. Additionally, while we appreciate the wildlife connectivity discussion in the response to comments, it still does little to nothing to assure conservation of or replacement for the foraging habitat of Swainson’s hawks that will be developed by the proposed project.
V. CONCLUSION
Thank you for the opportunity to submit comments on the Revised Final Environmental Impact Report for the Giovannoni Logistics Project. Because of the serious deficiencies in the document, we urge the City Council to delay approval so that the EIR can be revised and recirculated.
Given the possibility that the Center will be required to pursue legal remedies in order to ensure that the City complies with its legal obligations including those arising under CEQA, we would like to remind the County of its statutory duty to maintain and preserve all documents and communications that may constitute part of the “administrative record” of this proceeding. (§ 21167.6(e); Golden Door Properties, LLC v. Superior Court ((2020) 53 Cal.App.5th 733.) The administrative record encompasses any and all documents and communications that relate to any and all actions taken by the City with respect to the Project, and includes “pretty much everything that ever came near a proposed [project] or [] the agency’s compliance with CEQA . . . .” (County of Orange v. Superior Court (2003) 113 Cal.App.4th 1, 8.) The administrative record further includes all correspondence, emails, and text messages sent to or received by the City’s representatives or employees, that relate to the Project, including any correspondence, emails, and text messages sent between the City’s representatives or employees February 21, 2023
and the Applicant’s representatives or employees. Maintenance and preservation of the administrative record requires that, inter alia, the City (1) suspend all data destruction policies; and (2) preserve all relevant hardware unless an exact replica of each file is made.
Please do not hesitate to contact the Center with any questions at the number or email listed below.
Sincerely,
Frances Tinney
Legal Fellow
Center for Biological Diversity
1212 Broadway, Suite #800
Oakland, CA 94612
Tel: (509) 432-9256
[email protected]
February 21, 2023
References
Ambrose, Richard. 2000. “Wetland Mitigation in the United States: Assessing the Success of Mitigation Policies.” Wetlands (Australia) 19 (January). https://doi.org/10.31646/wa.242.
Ambrose, Richard, John C. Callaway, and Steven F. Lee. 2006. “An Evaluation of Compensatory Mitigation Projects under the Clean Water Act - Google Search.” https://www.waterboards.ca.gov/water_issues/programs/cwa401/docs/wetlandmitstudy_rpt.pdf.
Calhoun, Aram JK, J. Arrigoni, Malcolm L. Hunter Jr, and C. Richter. 2014. “Creating Successful Vernal Pools: A Literature Review and Advice for Practitioners.” Wetlands 34: 1027–38. https://scholar.archive.org/work/6xzvurqdrzbutkd5a5fhwrxlc4/access/wayback/https://digitalcom mons.library.umaine.edu/cgi/viewcontent.cgi?article=1031&context=mitchellcenter_pubs Estep J. A. Biology, movements, and habitat relationships of the Swainson’s Hawk in the Central Valley of California, 1986–87, California Department of Fish and Game, Nongame Bird and Mammal Section Report, 1989 Available at https://nrm.dfg.ca.gov/FileHandler.ashx?DocumentID=4019 Schlatter, Karen J., Akasha M. Faist, and Sharon K. Collinge. r2016. “Using Performance Standards to Guide Vernal Pool Restoration and Adaptive Management.”Restoration Ecology 24(2):145-152 https://www.researchgate.net/profile/AkashaFaist/publication/291553484_Using_performance_standards_to_guide_vernal_pool_restoration_a nd_adaptive_management/links/59ee6442a6fdcc32187db837/Using-performance-standards-toguide-vernal-pool-restoration-and-adaptive-management.pdf Sudol, Mark F., and Richard F. Ambrose. 2002. “The US Clean Water Act and Habitat Replacement:
Evaluation of Mitigation Sites in Orange County, California, USA.” Environmental Management 30 (5): 727–34. https://doi.org/10.1007/s00267-002-2787-3.
Woodbridge B, Finley KK, Bloom PH. 1995. Reproductive performance, age structure, and natal dispersal of Swainson’s Hawks in the Butte Valley, California. Journal of Raptor Research 29:187–192 Giovannoni Logistics Center Comments
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02/13/2023
Yvonne Baginski
[email protected]
7076945486
I own a home in American Canyon
I rent a home in American Canyon
I own a business in American Canyon
I work in American Canyon
None of the above
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Napa
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ca
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94558
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united states
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3205 Montclair Ave
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Please reconsider the Giovannoni Project in light of the need to provide a conservation easement for the foraging grasslands needed for the Swainson Hawk and Golden Eagle. If anything, the impact of the cumulative development in this area is crowding wildlife habitat and breaking them up in such a way that animal populations are unable to migrate, feed or roam freely. This is of significant concern.
The annihilation of 147.5 acres of foraging habitat needs more attention. This isn't just one piece of property standing on its own, it is also connected to a property area of south Napa County which is destined for environmental destruction where, without stopping it now, will mean an extinction of numerous foraging birds in south county. This Giovannoni project alone will wipe out 7 percent of the foraging habitat in the area. The other projects, Amazon, for example, have destroyed even more.
We need a sober look at this reality. We need to develop a strategy to lessen the impact of this project. We need to create a biologically diverse conservation easement for the remaining wildlife in south County. The Giovannoni Project would be a place to start. Please reconsider this effort to do the least harm to the environment. Among out suggestions, we include the consideration of a wildlife conservation easement which will be monitored by a credible manager, landscaping with only native plants, banning poison bait stations for the killing of rodents and other small mammals and demonstrating sustainable possibilities for all future buildings in the vicinity.