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The preponderance of short-term objectives and lack of systematic monitoring of restoration projects limits opportunities to learn from past experience and improve future restoration efforts. We conducted a retrospective, cross-sectional survey of 89 riparian revegetation sites and 13 nonrestored sites. We evaluated 36 restoration metrics at each site and used project age (0–39 years) to quantify plant community and aquatic habitat trajectories with a maximum likelihood model selection approach to compare linear and polynomial relationships. We found significant correlations with project age for 16 of 21 riparian vegetation, and 11 of 15 aquatic habitat attributes. Our results indicated improvements in multiple ecosystem services and watershed functions such as diversity, sedimentation, carbon sequestration, and available habitat. Ten riparian vegetation metrics, including native tree and exotic shrub density, increased nonlinearly with project age, while litter and native shrub density increased linearly. Species richness and cover of annual plants declined over time. Improvements in aquatic habitat metrics, such as increasing pool depth and decreasing bankfull width-to-depth ratio, indicated potentially improved anadromous fish habitats at restored sites. We hypothesize that certain instream metrics did not improve because of spatial and/or temporal limitations of riparian vegetation to affect aquatic habitat. Restoration managers should be prepared to maintain or enhance understory diversity by controlling exotic shrubs or planting shade-tolerant native species as much as 10 years after revegetation.
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Development of Vegetation and Aquatic Habitat in
Restored Riparian Sites of California’s North Coast
Michael S. Lennox,1,2David J. Lewis,1Randall D. Jackson,3John Harper,4Stephanie Larson,1
and Kenneth W. Tate5
The preponderance of short-term objectives and lack of
systematic monitoring of restoration projects limits oppor-
tunities to learn from past experience and improve future
restoration efforts. We conducted a retrospective, cross-
sectional survey of 89 riparian revegetation sites and 13
nonrestored sites. We evaluated 36 restoration metrics at
each site and used project age (039 years) to quantify
plant community and aquatic habitat trajectories with a
maximum likelihood model selection approach to com-
pare linear and polynomial relationships. We found sig-
nificant correlations with project age for 16 of 21 ripar-
ian vegetation, and 11 of 15 aquatic habitat attributes.
Our results indicated improvements in multiple ecosystem
services and watershed functions such as diversity, sed-
imentation, carbon sequestration, and available habitat.
Ten riparian vegetation metrics, including native tree and
exotic shrub density, increased nonlinearly with project
age, while litter and native shrub density increased lin-
early. Species richness and cover of annual plants declined
over time. Improvements in aquatic habitat metrics, such
as increasing pool depth and decreasing bankfull width-
to-depth ratio, indicated potentially improved anadromous
fish habitats at restored sites. We hypothesize that certain
instream metrics did not improve because of spatial and/or
temporal limitations of riparian vegetation to affect aquatic
habitat. Restoration managers should be prepared to main-
tain or enhance understory diversity by controlling exotic
shrubs or planting shade-tolerant native species as much
as 10 years after revegetation.
Key words: site-specific riparian revegetation, trajectory
analysis, restoration monitoring, regional assessment, post-
project appraisal.
Revegetation is a common tool to restore riparian areas for
many reasons, often by excluding livestock and/or planting
native trees. The number of river and stream restoration
projects in the United States has steadily increased since the
1980s from 100 to over 4,000 projects per year (Bernhardt
et al. 2005; Palmer et al. 2007). In California, over $2 billion
was spent on river restoration since 1980 with riparian
management the most common project type (Kondolf et al.
2007), but there has been limited systematic documentation of
project effectiveness to provide quality habitat and watershed
1University of California Cooperative Extension, Sonoma County, 133 Aviation
Boulevard, 109, Santa Rosa, CA 95403, U.S.A.
2Address correspondence to M. Lennox, email
3Department of Agronomy, University of Wisconsin-Madison, 1575 Linden Drive,
Madison, WI 53706-1597, U.S.A.
4University of California Cooperative Extension, Mendocino County, 890 N. Bush
Street, Ukiah, CA 95482, U.S.A.
5Department of Plant Sciences, University of California, Davis, Mail Stop 1, One
Shields Avenue, Davis, CA 95616-8780, U.S.A.
©2009 Society for Ecological Restoration International
doi: 10.1111/j.1526-100X.2009.00558.x
functions (Kondolf et al. 2007; Miller & Hobbs 2007; Palmer
et al. 2007).
Evaluation of previously restored sites has provided valu-
able feedback for understanding riparian habitat response
to various stream rehabilitation practices (Frissell & Nawa
1992; Opperman & Merenlender 2004; Tompkins & Kon-
dolf 2007). Numerous studies quantified riparian vegetation
recovery (Platts 1981; Kauffman et al. 1997; Opperman &
Merenlender 2000) and indirect recovery of aquatic habitat
has followed woody riparian vegetation establishment (Hupp
& Osterkamp 1996; Opperman & Merenlender 2004; Corenblit
et al. 2007). Restored project sites offer opportunities to learn
about resulting community structure and ecosystem processes
beyond static endpoints provided by reference sites (Parker
1997); however, long-term research over multiple decades
has been limited to case studies unable to quantify regional
variability or unintended consequences in a holistic evalua-
Some have used the amount of time since project imple-
mentation in various forms of trajectory analysis to provide
timelines for achieving specific objectives (Zedler & Callaway
1999; Golet et al. 2008). Watershed management carries the
Restoration Ecology 1
Developing California’s Vegetation and Aquatic Habitats
expectation that certain important societal objectives will be
achieved over time as a result of vegetation interacting with
physical processes (e.g., stochastic flood events transporting
sediment and pollutants). Examples of these objectives include
diversity (Hobbs 1993; Hupp & Osterkamp 1996), pollina-
tion (Kremen et al. 2004), sedimentation (Hupp & Osterkamp
1996; Corenblit et al. 2007), trophic dynamics (Baxter et al.
2005; Muotka & Syrjanen 2007), carbon storage (USDA 2000;
Bush 2008), nutrient cycling (Kauffman et al. 2004; Sheibley
et al. 2006; Bush 2008), water quality (Phillips 1989; Peterson
et al. 2001; Houlahan & Findlay 2004), infiltration (Kauff-
man et al. 2004), flood retention (Hupp & Osterkamp 1996;
Corenblit et al. 2007), available habitat (Dobkin et al. 1998;
Opperman & Merenlender 2004), and habitat use (Dobkin
et al. 1998; Golet et al. 2008). However, the trajectory analysis
has not been applied to watershed management in a holistic
approach using numerous attributes to assess the recovery of
multiple ecosystem services (Kremen 2005).
We conducted a retrospective, cross-sectional survey (i.e.,
chronosequence) of site-specific riparian revegetation projects
in three northern California coastal counties. Riparian veg-
etation and aquatic habitat response to stream rehabilitation
was quantified in a trajectory analysis using regression rela-
tionships with project age for 36 restoration metrics at 102
sites to provide a holistic regional evaluation of long-term
success over multiple decades. We used these trajectories to
infer changes in ecosystem services and watershed functions
(Black 1997) provided by riparian restoration.
Project Identification
Riparian revegetation sites were located in the mixed oak
woodland and annual grassland of California’s north coast. The
region has a Mediterranean climate with cool wet winters and
hot dry summers. However, this coastal region of California
is cooler with more moderate rainfall than most hardwood
rangelands. During the study period, mean annual precipitation
in the study area was 1,019 mm (range =679 1,629 mm)
and mean annual temperatures were 13.7C (range =12.0
15.1C). Streams and rivers in the region are dominated by
varying degrees of channel incision (Darby & Simon 1999)
and are located in watersheds with an average area of 23.5 km2
(range =0.2133.1km
2), elevation of 145.3 m asl (range =
3.7656.4 m asl), and 21.9% forested (range =0100%).
We surveyed 102 sites in Marin, Mendocino, and Sonoma
Counties (Fig. 1). Sites were selected in collaboration with
consultants, agencies, and landowners, whose permission was
solicited for access to conduct surveys. Project cooperators
identified both “successful” and “unsuccessful” projects to be
included in the study. Site selection focused on projects with
documented implementation dates in alluvial stream reaches
of willow and mixed oak woodland vegetation with few trees
present prior to project installation (e.g., Fig. 2a). Surveyed
project sites were primarily on second- and third-order streams
with a range in project age from 4 to 39 years since restoration.
Revegetation design at surveyed projects (n=89) was
site-specific and focused on establishing Salix species to
“jump start” recovery of riparian forests to control erosion
and sustain multiple watershed functions (Kauffman et al.
Figure 1. The three county study areas north of San Francisco Bay including locations of restored and nonrestored survey sites (image courtesy of
Sonoma County GIS Central) a). Aerial view of an idealized survey site depicting belt transects, plots delineated by landform, and herbaceous quadrats
b). Stream channel cross-section showing landforms along a transect c).
2Restoration Ecology
Developing California’s Vegetation and Aquatic Habitats
Figure 2. Photographic time-series of an example project site on a tributary to Walker Creek in Marin County, documenting vegetation response at 0 a),
2 b), 8 c), and 12 years d) since restoration occurred (images courtesy of Marin Resource Conservation District).
1997). The methods utilized were often implemented as
combinations of practices including tree or shrub planting
with dormant willow posts or container plants (Johnson
2003), biotechnical bank stabilization (Johnson 2003; Flosi
et al. 2004), and passive restoration (Kauffman et al. 1997)
using large herbivore management (e.g., removal, reduced
stocking rate, or exclusionary fencing for livestock and/or
deer). Nonrestored sites were surveyed (n=13) where local
experts indicated that a particular stream reach had vegetation
similar in structure to the project site before revegetation
Site Characterization
We characterized riparian forest and aquatic habitats at riparian
restoration project sites using 36 ecological attributes collected
at 5 nested spatial scales: (1) site (n=102, Fig. 1a), (2)
belt transect (n=3 per site, Fig. 1b), (3) landform class
(n=4 per transect, Fig. 1c), (4) plot (n=2+per landform),
and (5) quadrat (n=3 per plot, Fig. 1b). Landform classes
were delineated by channel morphology and depositional
or erosional features adapted from Harris (1987, 1999).
Specifically, we used the lowest observed bankfull location
and flood-prone elevation (2 ×bankfull depth) described by
Rosgen (1996) to delineate plots in the active floodplain. The
final plot sampled on each bank extended from the top of
the bank to the fence or field edge, and included alluvial
valley, terrace, or upland hillside geomorphic features. This
landform-based approach to collecting vegetation data allowed
for comparable results to be analyzed from various types of
stream channels.
At the site scale, data collected included small woody debris
(diameter <12 in), large woody debris (diameter >12 in),
and aggregate woody debris (debris jam clumps of 4 or more
pieces) counted within the bankfull channel (Flosi et al. 2004).
Pool characteristics assessed were mean pool depth,maximum
pool depth,pool frequency and percent pool habitat type (Flosi
et al. 2004). We collected stream substrate data at each site
and calculated percent fine sediment and embeddedness (Flosi
et al. 2004). The linear distance of riparian shade over the
thalweg was recorded at intervals with a hip chain as linear
channel canopy.
We placed three cross-sections and transects perpendicu-
lar to the channel stratified within each site at fast-water
riffle locations. Stream width and depth were measured and
documented as bankfull width-to-depth ratio (Rosgen 1996).
Streambank stability was assessed for both banks at each cross-
section according to Platts et al. (1987) and bank angle was
measured using a clinometer. Canopy density was measured
with a spherical densiometer following California Department
of Fish and Game protocols (Flosi et al. 2004) and solar radia-
tion was measured with a solar pathfinder by using the month
of August to standardize values before calculating intercepted
Restoration Ecology 3
Developing California’s Vegetation and Aquatic Habitats
solar radiation (Platts et al. 1987). Both measurements were
taken from the thalweg at each cross-section.
Data gathered within each plot included woody vegetation
density (trees >1m)andcanopy cover. Species identification
followed Hickman (1993). Herbaceous vegetation cover was
estimated using a modified Daubenmire Frame (20 ×50 cm)
to stratify quadrats equidistant in each plot perpendicular to
the stream channel (BLM 1996). The metric ground cover
included the sum of litter, vegetation, and stone cover (BLM
1996). Relative cover was calculated for six herbaceous
functional groups. Documenting survival was not possible
because of the lack of consistent record keeping on specific
numbers of plant species installed during the restoration
project and difficulty finding individual plantings in the field
at the oldest restored sites.
Data Analysis
We focused our analysis on detecting relationships between
project age and riparian forest and aquatic habitat metrics.
Plot and stream cross-section data were summarized into one
mean value of each metric by site for analysis to avoid
pseudoreplication (Hurlbert 1984). We then tested each metric
for curvilinear or linear fits. Models were constructed with
the generalized least squares function in S-Plus version 8
(Insightful Corp., Seattle, WA). Polynomial, linear, and null
(intercept only) models were compared with likelihood ratio
tests. If the models were significantly different (P <0.05),
we chose the model with the lowest Akaike Information
Criteria (AIC; Akaike 1974), otherwise the model with fewer
parameters was selected. If a linear model was better than the
polynomial model, we compared the linear model to a model
with no slope parameter using the same approach. Once best
fits were determined, the same parameters were estimated with
least squares regression to extract multiple R2values as an
assessment of goodness-of-fit.
Riparian Vegetation
Sixteen of 21 riparian vegetation metrics were significantly
related to project age, including 12 positive and 4 negative tra-
jectories (Table 1). The considerable increase over time in total
woody vegetation (Fig. 3), native tree, and exotic shrub/vine
densities were best characterized by polynomial relationships
with project age, but only total woody vegetation had a rel-
atively good fit. Exotic tree density did not demonstrate a
significant trajectory while the best fit for native shrub/vine
density was linear, but the fit was poor (Table 1).
Total canopy cover, native tree canopy cover, ground
cover, and exposed root cover increased curvilinearly as a
Table 1. Riparian vegetation parameter estimates for best fits determined by likelihood ratio tests (P <0.05) comparing polynomial, linear, and null
models using generalized least squares.
Parameter Estimates
Restoration Metric Best Fit y-intercept xx
Density (individuals ha1)
Total woody vegetation polynomial 459.8 329.9–7.60.39
Native tree polynomial 145.560.6–1.50.16
Native shrub/vine linear 204.825.1 0.08
Exotic tree n.s. 4.6—
Exotic shrub/vine polynomial 32.392.2–1.90.13
Absolute cover (%)
Total canopy polynomial 11.64.9–0.09 0.56
Native tree canopy polynomial 10.74.7–0.09 0.54
Ground cover polynomial 81.90.4–0.01 0.04
Exposed root polynomial 0.30.5–0.01 0.26
Total vegetation linear 43.2–0.3 0.05
Litter linear 19.90.4 0.21
Relative cover (%)
Native perennial grass n.s. 4.5—
Native perennial forb n.s. 2.5—
Exotic perennial grass n.s. 2.9—
Exotic perennial forb n.s. 1.8—
Annual grass polynomial 15.3–0.80.01 0.28
Annual forb polynomial 10.3–0.60.008 0.30
Species richness (spp. plot1)—
Tree polynomial 0.60.16 0.004 0.27
Shrub/vine polynomial 0.40.1–0.002 0.24
Perennial herbaceous polynomial 1.90.1–0.004 0.14
Annual herbaceous linear 4.4–0.1 0.21
Correlation coefficient (R2) determined with ordinary least squares regression.
4Restoration Ecology
Developing California’s Vegetation and Aquatic Habitats
Figure 3. Vegetation attributes as a function of project age (n=102) for
total woody density a), total canopy cover b), native tree cover c), and
annual forbs relative cover d).
function of project age, while litter cover increased in a linear
positive manner and total vegetation cover decreased linearly.
Native and exotic perennial grass and forb results were highly
variable and no significant relationships with project age were
found. Relative cover of annual forbs (Fig. 3) and grasses
had negative curvilinear trajectories. Species richness metrics
had positive curvilinear relationships to project age for the
tree, shrub/vine, and perennial herbaceous functional groups.
Annual species richness decreased linearly as project age
increased. Of all these significant relationships, the best fits
were total canopy cover and native tree canopy cover (Fig. 3).
Aquatic Habitat
Significant relationships with project age were observed for 11
of 15 aquatic habitat metrics, including eight positive and three
negative trajectories (Table 2). Stream channel morphology
results had significant trajectories for five of the six attributes.
The width-to-depth ratio of the bankfull channel had a negative
linear relationship with project age. Streambank stability had
a positive curvilinear relationship with project age and no
relationship was found for bank slope angle. The three woody
debris frequency metrics increased over time (Fig. 4). Small
and large wood frequencies were best described by curvilinear
relationships with project age, while aggregate debris jams of
wood were best described by a linear relationship.
Water column attributes had significant trajectories for six
of the nine investigated. Stream shade metrics, including
intercepted solar radiation, canopy density, and linear channel
canopy all increased curvilinearly over time (Fig. 4). Fine
sediment and embeddedness showed no significant trajectory.
Pool habitat metrics that had curvilinear relationships with
project age were maximum and mean pool depth as well as
pool habitat type. Pool frequency was not significantly related
to project age (Table 2).
Riparian Vegetation
While many significant polynomial and linear relationships
with project age were detected, most were relatively weak
as indicated by the R2values. However, we expected high
variability given the complex biophysical settings inherent
to riparian ecosystems specifically and Mediterranean climate
in general. The fact that we detected trajectories at all
indicates their broad application and importance to understand
fundamental changes following restoration.
Site-specific revegetation strategies accomplished the main
objectives of increasing woody species abundance and diver-
sity. Native tree establishment was the focus of revegeta-
tion efforts, so the large increases in tree density and cover
were expected (Fig. 2). Overall, an indirect plant community
response was predicted to follow a successional shift over
time from exotic annual herbaceous species to woody veg-
etation composed of overstory trees with a mosaic of native
shrubs and herbaceous perennials (Parker 1997; Dobkin et al.
1998). We detected this basic sequence, although native peren-
nial grasses and forbs did not show any long-term directional
trend, and shrubs colonized faster than has been observed at
more xeric inland riparian areas (Dobkin et al. 1998). Tree
density peaked 1525 years after restoration. Canopy cover
increase was relatively rapid indicating improved terrestrial
habitat for birds (Dobkin et al. 1998; White et al. 2005; Golet
et al. 2008), amphibians (USFWS 2002; Bulger et al. 2003),
and various wildlife species (Golet et al. 2008). In addition,
Restoration Ecology 5
Developing California’s Vegetation and Aquatic Habitats
Table 2. Aquatic habitat parameter estimates for best fits as determined by likelihood ratio tests (P <0.05) comparing polynomial, linear, and null models
using generalized least squares.
Parameter Estimates
Restoration Metric Best Fit y-intercept xx
Stream channel morphology
Bankfull width:depth ratio linear 35.5–0.6 0.10
Bank stability (%) polynomial 67.62.5–0.06 0.26
Bank slope (degrees) n.s. 15.2—
Small woody debris (count 100m–1)polynomial –0.40.5–0.007 0.48
Large woody debris (count 100m–1)polynomial –0.20.1–0.002 0.32
Aggregate woody debris (count 100m–1)linear 0.003 0.07 0.34
Water column
Intercepted solar radiation (%) polynomial 19.14.8–0.08 0.52
Canopy density (%) polynomial 12.35.0–0.08 0.49
Linear channel canopy (%) polynomial 0.55.8–0.10.49
Fine sediment (%) n.s. 15.3—
Embeddedness (%) n.s. 41.2—
Pool habitat (%) polynomial 28.71.8–0.04 0.10
Pool frequency (count 100m–1)n.s. 3.3—
Maximum pool depth (m) polynomial 0.60.04 0.0009 0.19
Mean pool depth (m) polynomial 0.40.03 0.0007 0.18
Correlation coefficient (R2)determined with ordinary least squares regression.
Figure 4. Aquatic habitat attributes as a function of project age (n=102) for small woody debris a), large woody debris b), aggregate woody debris
jams c), intercepted solar radiation d), canopy density e), and linear channel canopy f).
6Restoration Ecology
Developing California’s Vegetation and Aquatic Habitats
riparian vegetation changes at restored sites indicated improve-
ments in ecosystem services such as carbon storage via greater
tree abundance (USDA 2000). Other ecosystem services that
may be improved under these trajectories include diversity
(Hupp & Osterkamp 1996; Hobbs 1993), pollination (Kremen
et al. 2004), sedimentation (Hupp & Osterkamp 1996; Coren-
blit et al. 2007), nutrient cycling (Peterson et al. 2001; Kauff-
man et al. 2004; Sheibley et al. 2006), and trophic dynamics
(Baxter et al. 2005; Muotka & Syrjanen 2007).
The increase in exotic shrub density over time was unin-
tended and undesirable. This phenomenon has been noted
in past work (Borgmann & Rodewald 2005; Badano et al.
2007). Exotic tree abundance did not correlate with project
age, but these taxa were occasionally present at restored
sites from previous plantings. In contrast, the most common
exotic shrub, Himalayan blackberry (Rubus discolor ), domi-
nated many older restored sites (greater than 20 years old) by
establishing homogeneous patches, which is similar to obser-
vations by Lambrecht-McDowell and Radosevich (2005). The
rapid trajectory of exotic shrub abundance reduces options
for management in the riparian corridor. Consideration of
exotic vegetation should focus on the trade-offs that exotic
species present for achieving management goals over multi-
ple decades (Parker 1997). For example, White et al. (2005)
found juvenile Swainson’s Thrush (Catharus ustulatus)used
Himalayan blackberry for cover and food, so removing this
vegetation from recently restored sites may affect wildlife pop-
ulations negatively. However, delaying active control of exotic
shrubs past the initial 20 years of restoration may eliminate
chances for adaptive management and cost effective solutions,
as explained by Zavaleta (2000).
It was not surprising that perennial herbaceous species did
not respond to restoration since the focus of revegetation
was woody species. Holl and Crone (2004) made similar
observations. Annual vegetation was clearly reduced over
time, but resurgence of native perennial grasses and forbs is not
likely without significant propagule supply (Bartolome et al.
2004) from flood inundation (Hupp & Osterkamp 1996) and
less competition from exotic (Holl & Crone 2004) or shrub
species (Brown & Archer 1999).
Aquatic Habitat
A primary purpose for establishing native trees, in particular
Salix species, was to stabilize streambanks (Johnson 2003)
because forested vegetation contains the greatest fine root
density for erosion resistance (Wynn et al. 2004) and tree den-
sity increases channel roughness increasing sedimentation and
retention of flood water (Hupp & Osterkamp 1996; Coren-
blit et al. 2007). Therefore, the changes we found in stream
channel morphology and streambank stability were expected
and should result in improved water quality with less chronic
sediment delivery to streams from restored sites (NCRWQCB
1998; Corenblit et al. 2007). Decreasing the bankfull chan-
nel width-to-depth ratio was also an expected response from
revegetation because stream channels tend to deepen and nar-
row as sedimentation on floodplains increases following tree
establishment (Hupp & Osterkamp 1996; Opperman & Meren-
lender 2004; Corenblit et al. 2007). This process was enhanced
by live wood interacting with woody debris forming persistent
instream structure, as explained by Opperman and Merenlen-
der (2007). The accumulation of large wood and debris jams
provides greater complexity of instream habitat such as deeper
pools (Beechie & Sibley 1997) and cover (Cederholm et al.
Improved pool habitat and depth indicate greater abundance
and diversity of aquatic fauna may be able to use habitat
at restored sites as complexity within the water column
increased over time. Pools provide cover that protect prey from
predators, create slower flow niches during winter storms, and
contribute to temperature stratification for thermal refugia in
summer (Ebersole et al. 2001). The large increase of stream
shade attributes over time was an expected outcome and
indicates water temperature may be reduced following riparian
revegetation (Brown 1969; Opperman & Merenlender 2004).
Aquatic habitat metrics that did not improve over time offer
further insight into biogeomorphic processes in the riparian
zone (Corenblit et al. 2007). Fine sediment and embeddedness
of stream channel substrate did not change indicating that
these metrics may be linked to watershed processes operating
at spatial scales larger than those of the typical revegetation
project site (Houlahan & Findlay 2004; Opperman et al. 2005).
Moreover, the temporal range of our survey may not have been
sufficient to encompass change in these parameters.
While long-term monitoring of individual sites would have
produced a clearer understanding of riparian vegetation and
aquatic habitat trajectories following restoration, the sub-
stitution of space-for-time in our chronosequence compar-
isons provided useful insights that inform regional restoration
efforts. This cross-sectional survey approach also offers an
effective option for systematic, objective assessment of com-
pleted projects and postproject appraisals (Kondolf et al. 2007;
Tompkins & Kondolf 2007). We suggest that stream restoration
research further investigate the impact of establishing woody
species on stream channel morphology, nutrient cycling, and
overall plant diversity. This will prepare the restoration part-
nership to manage numerous objectives and ecosystem ser-
vices over multiple decades.
Implications for Practice
Site-specific riparian revegetation strategies were suc-
cessful in maintaining native tree and shrub density,
cover, and richness over multiple decades.
Shrub control may be important for maintaining under-
story diversity at restored riparian sites, since the trajec-
tory for exotic shrub abundance and variability in native
herbaceous species indicated a need for vegetation man-
agement 1020 years postrestoration.
Although aquatic habitat improved following revegeta-
tion (e.g., more shade, more woody debris, and deeper
pools), other important instream attributes such as fines
Restoration Ecology 7
Developing California’s Vegetation and Aquatic Habitats
and embeddedness did not recover over multiple decades
and may be controlled by watershed factors.
Monitoring of riparian revegetation projects should
include bank stability, woody debris, channel width-to-
depth ratio, and pool depth where appropriate, in addition
to plant diversity and cover over time.
We are grateful to the cooperative group of natural resource
managers whose contributions are the reason the project was
possible. We want to recognize the assistance provided by
Thomas Schott, Paul Sheffer, Gale Ranch, Paul Martin, Jeff
Opperman, and Lisa Bush as well as the Marin County
Resource Conservation District (RCD), Mendocino County
RCD, Southern Sonoma RCD, Natural Resources Conserva-
tion Service, Prunuske Chatham, Inc., Bay Institute’s Students
& Teachers Restoring A Watershed (STRAW), Bioengineering
Associates, and many more. Funding was provided by Califor-
nia Coastal Conservancy, National Oceanographic and Atmo-
spheric Administration’s Restoration Center, and University of
California Division of Agriculture and Natural Resources.
Akaike, H. 1974. A new look at the statistical model identification IEEE
Transactions on Automatic Control AC 19:716 723.
Badano, E. I., E. Villarroel, R. O. Bustamante, P. A. Marquet, and L. A.
Cavieres. 2007. Ecosystem engineering facilitates invasions by exotic
plants in high-Andean ecosystems. Journal of Ecology 95:682–688.
Bartolome, J. W., J. S. Fehmi, R. D. Jackson, and B. Alen-Diaz. 2004.
Response of a native perennial grass stand to disturbance in California’s
coast range grassland. Restoration Ecology 12:279 289.
Baxter, C. V., K. D. Fausch, and W. C. Saunders. 2005. Tangled webs:
Reciprocal flows of invertebrate prey link streams and riparian zones.
Freshwater Biology 50:201–220.
Beechie, T. J., and T. H. Sibley. 1997. Relationships between channel
characteristics, woody debris, and fish habitat in northwestern Washington
streams. Transactions of the American Fisheries Society 126:217– 219.
Bernhardt, E. S., M. A. Palmer, J. D. Allan, G. Alexander, K. Barnas,
S. Brooks, et al. 2005. Synthesizing U.S. river restoration efforts.
Science 308:636– 637.
Black, P. E. 1997. Watershed functions. Journal of the American Water
Resources Association 33:1–11.
BLM (Bureau of Land Management). 1996. Sampling vegetation attributes.
BLM National Applied Resources Center. BLM Technical Reference.
4400– 4404. Denver, Colorado.
Borgmann, K. L., and A. D. Rodewald. 2005. Forest restoration in urbanizing
landscapes: interactions between land use and exotic shrubs. Restoration
Ecology 13:334– 340.
Brown, G. W. 1969. Predicting temperatures of small streams. Water Resources
Research 5:68– 75.
Brown, J. R., and S. Archer. 1999. Shrub invasion of grassland: recruitment is
continuous and not regulated by herbaceous biomass or density. Ecology
80:2385– 2396.
Bush, J. K. 2008. Soil nitrogen and carbon after twenty years of riparian forest
development. Soil Science Society of America Journal 72:815 822.
Bulger, J. B., N. J. Scott, and R. B. Seymour. 2003. Terrestrial activity
and conservation of adult California red-legged frogs Rana draytonii in
coastal forests and grasslands. Biological Conservation 110:85– 95.
Cederholm, C. J., R. E. Bilby, P. A. Bisson, T. W Bumstead, B. R. Fransen,
W. J. Scarlett, and J. W. Ward. 1997. Response of juvenile coho salmon
and steelhead to placement of large woody debris in a coastal Washington
stream. North American Journal of Fisheries Management 17:947 963.
Corenblit, D., E. Tabacchi, J. Steiger, and A. M. Gurnell. 2007. Reciprocal
interactions and adjustments between fluvial landforms and vegetation
dynamics in river corridors: a review of complementary approaches.
Earth-Science Reviews 84:56 –86.
Darby, S. E., and A. Simon, editors. 1999. Incised river channels. John Wiley
& Sons, Ltd., West Sussex, England.
Dobkin, D. S., A. C. Rich, and W. H. Pyles. 1998. Habitat and avifaunal
recovery from livestock grazing in a riparian meadow system of the
northwest Great Basin. Conservation Biology 12:209 221.
Ebersole, J. L., W. J. Liss, and C. A. Frissel. 2001. Relationship between
stream temperature, thermal refugia and rainbow trout Onchorhynchus
mykiss abundance in arid-land streams in the northwestern United States.
Ecology of Freshwater Fish 10:1 10.
Flosi, G., S. Downie, J. Hopelain, M. Bird, R. Coey, and B. Collins.
2004. California salmonid stream habitat restoration manual. California
Department of Fish and Game. 3rd edition. Inland Fisheries Division,
Sacramento, California.
Frissell, C. A., and R. K. Nawa. 1992. Incidences and causes of physical
failure of artificial habitat structures in stream of western Oregon
and Washington. North American Journal of Fisheries Management
12:182– 197.
Golet, G. H., T. Gardali, C. A. Howell, J. Hunt, R. A. Luster, W. Rainey,
M. D. Roberts, J. Silveira, H. Swagerty, and N. Williams. 2008. Wildlife
response to riparian restoration on the Sacramento river. San Francisco
Estuary and Watershed Science 6:1–26.
Harris, R. R. 1987. Occurrence of vegetation on geomorphic surfaces in
the active floodplain of a California alluvial stream. American Midland
Naturalist 118:393– 405.
Harris, R. R. 1999. Defining reference conditions for restoration of riparian
plant communities: examples from California, USA. Environmental
Management 24:55– 63.
Hickman, fnmJ. C., editor. 1993. The Jepson manual to higher plants of
California. University of California Press, Berkeley.
Hobbs, R. J. 1993. Can revegetation assist in the conservation of biodiversity
in agricultural landscapes? Pacific Conservation Biology 1:29–38.
Holl, K. D., and E. E. Crone. 2004. Applicability of landscape and island
biogeography theory to restoration of riparian understorey plants. Journal
of Applied Ecology 41:922–933.
Houlahan, J. E., and C. S. Findlay. 2004. Estimating the critical distance at
which adjacent land-use degrades wetland water and sediment quality.
Landscape Ecology 19:677–690.
Hupp, C., and W. Osterkamp. 1996. Riparian vegetation and fluvial geomor-
phic processes. Geomorphology 14:277 295.
Hurlbert, S. H. 1984. Pseudoreplication and the design of ecological field
experiments. Ecological Monographs 54:187 –211.
Johnson, C. 2003. Five low-cost methods for slowing streambank erosion.
Journal of Soil and Water Conservation 58:12– 17.
Kauffman, J. B., R. L. Beschta, N. Otting, and D. Lytjen. 1997. An ecological
perspective of riparian and stream restoration in the Western United
States. Fisheries 22:12–24.
Kauffman, J. B., A. S. Thorpe, and E. N. J. Brookshire. 2004. Livestock
exclusion and below ground ecosystem response in riparian meadows of
eastern Oregon. Ecological Applications 14:1671 1679.
Kondolf, G. M., S. Anderson, R. Lave, L. Pagano, A. Merenlender, and
E. S. Bernhardt. 2007. Two decades of river restoration in California:
what can we learn? Restoration Ecology 15:516 523.
Kremen, C., N. M. Williams, R. L., Bugg, J. P. Fay, and R. W. Thorp. 2004.
The area requirements of an ecosystem service: crop pollination by native
bee communities in California. Ecology Letters 7:1109–1119.
8Restoration Ecology
Developing California’s Vegetation and Aquatic Habitats
Kremen, C. 2005. Managing ecosystem services: what do we need to know
about their ecology? Ecology Letters 8:468 479.
Lambrecht-McDowell, S. C., and S. R. Radosevich. 2005. Population demo-
graphics and tradeoffs to reproduction of an invasive and noninvasive
species of Rubus. Biological Invasions 7:281–295.
Miller, J. R., and R. J. Hobbs. 2007. Habitat restoration-do we know what
we’re doing? Restoration Ecology 15:282 390.
Muotka, T., and J. Syrjanen. 2007. Changes in habitat structure, benthic
macroinvertebrate diversity, trout populations and ecosystem processes
in restored forest streams: a boreal perspective. Freshwater Biology
52:724– 737.
NCRWQCB (North Coast Regional Water Quality Control Board). 1998.
Water-quality attainment strategy (total maximum daily load) for sedi-
ment for the Garcia River watershed. Santa Rosa, California.
Opperman, J., and A. Merenlender. 2000. Deer herbivory as an ecological
constraint to restoration of degraded riparian corridors. Restoration
Ecology 8:41– 47.
Opperman, J. J., and A. M. Merenlender. 2004. The effectiveness of
riparian restoration for improving instream fish habitat in four hardwood-
dominated California streams. North American Journal of Fisheries
Management 24:822– 834.
Opperman, J. J., K. O. Lohse, C. Brooks, N. M. Kelly, and A. M. Merenlender.
2005. Influence of land use on fine sediment in salmonid spawning
gravels within the Russian River Basin, California. Canadian Journal of
Fishery Aquatic Science 62:2740– 2751.
Opperman, J. J., and A. M. Merenlender. 2007. Living trees provide
stable large wood in streams. Earth Surface Processes and Landforms
32:1229– 1238.
Palmer, M., J. D. Allan, J. Meyer, and E. S. Bernhardt. 2007. River restoration
in the twenty-first century: data and experiential knowledge to inform
future efforts. Restoration Ecology 15:472 481.
Parker, V. T. 1997. The scale of successional models and restoration objectives.
Restoration Ecology 5:301 –306.
Peterson, B. J., W. M. Wollheim, P. J. Mulholland, J. R. Webster, J. L. Meyer,
J. L. Tank, et al. 2001. Control of nitrogen export from watersheds by
headwater streams. Science 292:86– 90.
Phillips, J. D. 1989. Nonpoint source pollution control effectiveness of riparian
forests along a coastal plain river. Journal of Hydrology 110:221 238.
Platts, W. S. 1981. Sheep and streams. Rangelands 3:158–160.
Platts W. S., C. Armour, G. D. Booth, M. Bryant, J. L. Bufford, P. Culpin,
et al. 1987. Methods for evaluating riparian habitats with applications to
management. USDA Forest Service. General Technical Report INT-221.
Ogden, Utah. 177 p.
Rosgen, D. 1996. Applied river morphology. Wildland Hydrology, Pagosa
Springs, Colorado.
Sheibley, R. W., D. S. Ahearn, and R. A. Dahlgren. 2006. Nitrate loss
from a restored floodplain in the lower Cosumnes River, California.
Hydrobiologia 571:261–272.
Tompkins, M. R., and G. M. Kondolf. 2007. Systematic postproject appraisals
to maximize lessons learned from river restoration projects: case study of
compound channel restoration projects in northern California. Restoration
Ecology 15:524– 537.
USDA (U.S. Department of Agriculture). 2000. Growing carbon: a new crop
that helps agricultural producers and the climate too. Natural Resources
Conservation Service, Washington, DC.
USFWS (U.S. Fish and Wildlife Service). 2002. Recovery plan for the
California red-legged frog (Rana aurora draytonii). Pages viii +173.
U.S. Fish and Wildlife Service, Portland, Oregon.
White, J. D., T. Gardali, F. R. Thompson, and J. Faaborg. 2005. Resource
selection by juvenile Swainson’s Thrushes during the postfledging period.
The Condor 107:388– 401.
Wynn, T. M., S. Mostaghimi, J. A. Burger, A. A. Harpold, M. B. Henderson,
and L. A. Henry. 2004. Ecosystem restoration: variation in root density
along stream banks. Journal of Environmental Quality 33:2030– 2039.
Zavaleta, E. 2000. The economic value of controlling an invasive shrub. Ambio
29:462– 467.
Zedler, J. B., and J. C. Callaway. 1999. Tracking wetland trajectory: do
mitigation sites follow desired trajectories? Restoration Ecology 7:69–73.
Restoration Ecology 9
... It is indicated that the evaluation of a better-known taxonomic group may be inadequate to represent the response of biodiversity in the restored area. Apart from biodiversity, the abundance of understory plants, increase in native plants and herbaceous flora, decrease in exotic or weed plants in the restored area, etc., are all considered for the assessment (McClain et al., 2011;Lennox et al. 2011;Hough-Snee et al. 2013). According to Salinas et al. (2000), simple planting and monitoring of plants after seven years indicated the success of the restoration. ...
... The abiotic factors can be generally categorised into stream-related and others (Fig. 5). Stream-related measurements include analysis of channel morphology, suspended sediment content, and sedimentation process (Lennox et al. 2011). De Mello et al. (2017 used watershed simulation modeling to simulate streamflow, suspended sediments and nutrients. ...
... Temporal variability in climate such as an increase in precipitation, temperature variations shall be considered for understanding the success of riparian restoration (Meyer et al. 2010). Lennox et al. (2011) confirmed the availability of habitat to be a significant indicator of ecosystem service by the restored riparian ecosystem, as seen from the potentially improved anadromous fish habitats at the restored site. ...
Ecosystems across the globe, be it terrestrial, marine or transitional in nature are under pressure due to multiple drivers of changes including anthropogenic. Restoring the vitality of degraded systems is crucial for fulfilling the UN-Sustainable Development Goals in a timely manner. It is also essential for attaining the targets of the ambitious UN-Decade on Ecosystem Restoration (UN-DER). Riparian ecosystems are one among systems undergoing drastic changes due to anthropogenic pressures. They are a heterogeneous and biodiversity rich system due to its transitional zone occurrence between terrestrial and aquatic realms, including riverbanks, floodplains and wetlands, and provide ecosystem services on both local as well as global levels. Here we review the prospects of restoring riparian ecosystems in the context of the UNDER. Even though the momentum for restoring riparian habitats began in the 1970s, our study reveals that intensive restoration programmes across the world are sparse and more efforts are needed to restore degraded riparian systems for regaining ecosystem health and complexity. Furthermore, an in-depth analysis of various strategies deployed for restoring riparian ecosystems around the world reveals that a participatory approach and site-specific strategies are needed for better output. Also, active along with passive restoration is required for better recovery. We suggest a three-stage strategy-preassessment, restoration activities and post monitoring and maintenance. It includes the involvement of stakeholders across all stages, which also supports their livelihoods. The restoration of riparian ecosystems supports the targets of UNDER while providing both global as well as local ecosystem services.
... Sin embargo, el cese de las actuaciones de mantenimiento causó su disminución en el periodo 2019-2021, siendo la tendencia a situarse en valores cercanos a los iniciales de 2015. Otros estudios (Lennox et al., 2011) confirman que la riqueza disminuye a medida que la antigüedad del proyecto avanza. Esto se debe posiblemente a que las especies nativas plantadas requieren más tiempo para desarrollarse y competir con A. donax, que ocupa progresivamente el espacio ribereño e intercepta la luz solar (Jiménez-Ruiz & Santín-Montanyá, 2016). ...
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Las riberas fluviales constituyen áreas de gran interés ecológico, pero su modificación antrópica ha provocado el establecimiento de especies invasoras, como Arundo donax L. en la cuenca del río Segura (SE de España), con efectos perjudiciales sobre la biodiversidad. Ante esta problemática, en el marco del proyecto LIFE+ RIPISILVANATURA (2015-2019) se ejecutaron actuaciones de control de la caña y plantaciones con especies nativas en el tramo medio del río Segura. En este trabajo se realiza un seguimiento a medio plazo tras dos años de la finalización del proyecto y cese o relajación de las tareas de mantenimiento (desbroces de caña) en 8 de las estaciones restauradas de cara a evaluar el grado de éxito alcanzado en comparación a las condiciones de referencia. Para detectar diferencias temporales en la composición y abundancia de especies en estaciones revegetadas y de referencia, se aplicaron análisis de escalamiento multidimensional no métrico (NMDS) y modelos lineales generalizados para datos de abundancia (GLM). Para determinar si la altura media de las especies plantadas difiere entre estaciones con o sin mantenimiento, se aplicaron ANOVAs. Además, se determinaron los cambios temporales en diferentes indicadores ecológicos y el efecto de la continuación o no del mantenimiento mediante modelos lineales de efectos mixtos. Tras 6 años del inicio y 2 años de la finalización del proyecto, el éxito alcanzado ha sido relativamente bajo. Aunque actualmente la composición de la vegetación de las estaciones restauradas tiende a ser más similar a las comunidades de referencia, la ausencia de mantenimiento ha producido en los últimos años un aumento de la cobertura, densidad y altura de A. donax, así como una disminución de la riqueza de especies y un estado ecológico de las riberas deficiente. Por tanto, un programa de mantenimiento y seguimiento a medio-largo plazo es fundamental para mejorar el éxito en futuros proyectos de restauración de riberas.
... Planning for restoration and achieving desired goals requires understanding the abiotic and biotic factors that shape the development of the natural communities that projects intend to recreate (Harris 1999). A key aspect of this process is predicting the outcome of revegetation efforts, and specifically, how plant community cover and structure develop over time (Lennox et al. 2011). In California, United States, riparian restoration is regulated by a state-level Lake and Streambed Alteration Agreement process that establishes performance criteria for vegetation survival and cover that must be met by the end of the project period (generally 5 years; California Department of Fish and Wildlife 2013). ...
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The investment to restore riparian systems necessitates employing data driven planning, implementation, and management. To evaluate the demography/growth of riparian plants established vegetatively, we measured three years of survival and biometrics for five tree/shrub species experimentally planted as pole cuttings within a large restoration program in a Mediterranean-climate riparian system. During the experiment, a recently introduced ambrosia beetle, Euwallacea whitfordiodendrus (polyphagous shot hole borer), colonized pole cuttings, providing an opportunity to evaluate beetle establishment rates among species. Pole cutting survival varied significantly among species and reflected their relative ability to propagate vegetatively. Baccharis salicifolia had over 90% survival, while Salix lasiolepis and S. laevigata had intermediate, and Populus trichocarpa and P. fremontii had low survival. Canopy growth reflected species life histories. The small tree/shrub, Salix lasiolepis had robust early growth, while slower growing, larger tree species (Salix laevigata, Populus trichocarpa, and P. fremontii) filled in canopy later. Only S. lasiolepis was colonized by E. whitfordiodendrus during the second year and had the greatest infestation rates in all years. All species except B. salicifolia were infested by the third year. Beetles tended to colonize larger trees. Our observed survival and growth trends can serve as a guide for estimating planting density and vegetation structural development over time, and to inform adaptive management of riparian species of Mediterranean-climate and other arid-land systems. Ecosystem stressors, such as invasive insects and their impacts to plant growth, must be considered during restoration as introductions are anticipated to continue with increased severity with climate change. This article is protected by copyright. All rights reserved.
... If non-native species are being introduced, it is important to assess and mitigate the associated risks (Sáenz-Romero et al., 2016;Simler et al., 2019;Weeks et al., 2011). Even when restoration involves only native species, there can be trade-offs that need to be managed to enhance biodiversity outcomes, as some species may benefit at the expense of others (Biel et al., 2017;Porensky et al., 2014), or species abundance could increase at the expense of species richness (Lennox et al., 2011). ...
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Nature‐based solutions (NbS)—solutions to societal challenges that involve working with nature—have recently gained popularity as an integrated approach that can address climate change and biodiversity loss, while supporting sustainable development. Although well‐designed NbS can deliver multiple benefits for people and nature, much of the recent limelight has been on tree planting for carbon sequestration. There are serious concerns that this is distracting from the need to rapidly phase out use of fossil fuels and protect existing intact ecosystems. There are also concerns that the expansion of forestry framed as a climate change mitigation solution is coming at the cost of carbon rich and biodiverse native ecosystems and local resource rights. Here, we discuss the promise and pitfalls of the NbS framing and its current political traction, and we present recommendations on how to get the message right. We urge policymakers, practitioners and researchers to consider the synergies and trade‐offs associated with NbS and to follow four guiding principles to enable NbS to provide sustainable benefits to society: (1) NbS are not a substitute for the rapid phase out of fossil fuels; (2) NbS involve a wide range of ecosystems on land and in the sea, not just forests; (3) NbS are implemented with the full engagement and consent of Indigenous Peoples and local communities in a way that respects their cultural and ecological rights; and (4) NbS should be explicitly designed to provide measurable benefits for biodiversity. Only by following these guidelines will we design robust and resilient NbS that address the urgent challenges of climate change and biodiversity loss, sustaining nature and people together, now and into the future.
... Mixed ecological outcomes occurred when intervention impacts differed across space (e.g., due to the displacement of drivers of degradation or habitat loss, exacerbating the pressure to surrounding locations; Mekuria et al., 2015), or when the intervention had a positive effect on some ecological attributes but not others (e.g., the intervention increased native species richness as well as increasing abundance of exotic species; Lennox et al., 2011). ...
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Nature-based solutions (NbS) to climate change currently have considerable political traction. However, national intentions to deploy NbS have yet to be fully translated into evidence-based targets and action on the ground. To enable NbS policy and practice to be better informed by science, we produced the first global systematic map of evidence on the effectiveness of nature-based interventions for addressing the impacts of climate change and hydrometeorological hazards on people. Most of the interventions in natural or semi-natural ecosystems were reported to have ameliorated adverse climate impacts. Conversely, interventions involving created ecosystems (e.g., afforestation) were associated with trade-offs; such studies primarily reported reduced soil erosion or increased vegetation cover but lower water availability, although this evidence was geographically restricted. Overall, studies reported more synergies than trade-offs between reduced climate impacts and broader ecological, social, and climate change mitigation outcomes. In addition, nature-based interventions were most often shown to be as effective or more so than alternative interventions for addressing climate impacts. However, there were substantial gaps in the evidence base. Notably, there were few studies of the cost-effectiveness of interventions compared to alternatives and few integrated assessments considering broader social and ecological outcomes. There was also a bias in evidence toward the Global North, despite communities in the Global South being generally more vulnerable to climate impacts. To build resilience to climate change worldwide, it is imperative that we protect and harness the benefits that nature can provide, which can only be done effectively if informed by a strengthened evidence base.
... Revegetation of riparian zones to achieve erosion control had been attempted on 32 of these reaches, and we included an additional 10 unrestored reaches for comparison. Revegetation projects at these sites were typically installed using a combination approach of three active intervention measures [29], including tree and shrub planting, biotechnical streambank stabilization (e.g., armoring with coarse woody debris), and large herbivore management (e.g., fencing to exclude livestock and/or deer, reduced stocking rate, or removal of grazing). For an analysis focused on the effects of restoration success and time since restoration, we chose riparian revegetation projects ranging in age from 1 to 45 years post-restoration, both successes and failures, that shared a high degree of similarity in geomorphology and landscape setting. ...
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Background: Globally, vegetation in riparian zones is frequently the target of restoration efforts because of its importance in reducing the input of eroded sediment and agricultural nutrient runoff to surface waters. Here we examine the potential of riparian zone restoration to enhance carbon sequestration. We measured soil and woody biomass carbon stocks, as well as soil carbon properties, in a long-term chronosequence of 42 streambank revegetation projects in northern California rangelands, varying in restoration age from 1 to 45 years old. Results: Where revegetation was successful, we found that soil carbon measured to 50 cm depth increased at a rate of 0.87 Mg C ha-1 year-1 on the floodplain and 1.12 Mg C ha-1 year-1 on the upper bank landform. Restored sites also exhibited trends toward increased soil carbon permanence, including an increased C:N ratio and lower fulvic acid: humic acid ratio. Tree and shrub carbon in restored sites was modeled to achieve a 50-year maximum of 187.5 Mg C ha-1 in the channel, 279.3 Mg ha-1 in the floodplain, and 238.66 Mg ha-1 on the upper bank. After 20 years of restoration, the value of this carbon at current per-ton C prices would amount to $US 15,000 per km of restored stream. Conclusion: We conclude that revegetating rangeland streambanks for erosion control has a substantial additional benefit of mitigating global climate change, and should be considered in carbon accounting and any associated financial compensation mechanisms.
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In the actual context of global change and biodiversity depletion, soil bioengineering represents an important tool for riparian ecosystem restoration and species conservation. Various techniques have already been implemented, but their adaptation still must be carried out in Caribbean Islands biodiversity hotspots, where suitable species remains unknown. Nitrogen-fixing legumes are particularly relevant for ecological restoration and the diversity of native Caribbean legume trees is promising in the search for suitable species for soil bioengineering. We hypothesized that Caribbean legume tree species present a growth performance and set of biotechnical traits compatible with their use in soil bioengineering. We selected five native legume trees, adapted to riparian environments, in different ecosystems (swamp forest, evergreen seasonal forest, rainforest) based on their ecology, resistance to disturbance and seed production characteristics. We measured root traits relevant for soil bioengineering on nursery grown 3-month-old seedlings. Despite their differences in sensitivity to herbivory and in growth strategies, the selected species have a high potential for use in soil bioengineering, with high seed production, high germination rates—from 88 to 100%—, and 100% survival rates, and are therefore compatible with large scale plant material production. We provided practical guidance tools for their integration into soil bioengineering techniques.
Livestock exclusion is a widespread restoration technique in the Pacific Northwest to protect and improve riparian and stream habitats. To assess stream restoration outcomes from excluding livestock, the Washington State Salmon Recovery Funding Board and the Oregon Watershed Enhancement Board evaluated 12 livestock exclusion projects from 2004 to 2017 using a before-after control-impact design. Paired treatment and control reaches were monitored once before restoration implementation (year 0) and several years after implementation (years 1, 3, 5, and 10) to assess bank erosion, bank canopy cover, riparian vegetation structure, pool tail fine sediment, and exclosure fencing function. Livestock exclusion significantly reduced bank erosion and bare ground. Bank erosion in treatment reaches decreased from 44% pre-project to 11% by year 10. In treatment reaches, bare ground was over 1.5 times lower in year 10 than pre-project. Most treatment reaches had intact fencing at the conclusion of the ten-year monitoring period, but there were instances where fencing did not fully function as intended, allowing livestock to access riparian areas inside the exclosure. Several metrics did not respond over time, which may be the result of several factors, including limitations of the sampling protocols, evidence of livestock grazing in treatment reaches, lack of site stratification, control reaches that were not well matched with treatment reaches, and short-duration of pre-project data collection. Despite these limitations, we still detected significant decreases in bank erosion and bare ground within treatment reaches. Future livestock exclusion monitoring should consider focus on ensuring fence maintenance, improved monitoring oversight, and the use of more quantitative monitoring protocols.
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Nature-based solutions (NbS) have been positioned and implemented in urban areas as solutions for enhancing urban resilience in the face of a wide range of urban challenges. However, there is a lack of recommendations of optimal NbS and appropriate typologies fitting to different contexts and urban design. The analytical frameworks for NbS implementation and impact evaluation, that integrate NbS into local policy frameworks, socio-economic transition pathways, and spatial planning, remain fragmented. In this article, the NbS concept and its related terminologies are first discussed. Second, the types of NbS implemented in Europe are reviewed and their benefits over time are explored, prior to categorizing them and highlighting the key methods, criteria, and indicators to identify and assess the NbS’s impacts, co-benefits, and trade-offs. The latter involved a review of the websites of 52 projects and some relevant publications funded by EU Research and Innovation programs and other relevant publications. The results show that there is a shared understanding that the NbS concept encompasses benefits of restoration and rehabilitation of ecosystems, carbon neutrality, improved environmental quality, health and well-being, and evidence for such benefits. This study also shows that most NbS-related projects and activities in Europe use hybrid approaches, with NbS typically developed, tested, or implemented to target specific types of environmental–social–economic challenges. The results of this study indicate that NbS as a holistic concept would be beneficial in the context of climate action and sustainable solutions to enhance ecosystem resilience and adaptive capacity within cities. As such, this article provides a snapshot of the role of NbS in urban sustainability development, a guide to the state-of-the-art, and key messages and recommendations of this rapidly emerging and evolving field.
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The concept of ecosystem services (ES) has risen to prominence based on its promise to vastly improve environmental decision‐making and to represent nature's many benefits to people. Yet the field has continued to be plagued by fundamental concerns, leading some to believe that the field of ES must mature or be replaced. In this paper, we quantitatively survey a stratified random sample of more than 1,000 articles addressing ES across three decades of scholarship. Our purpose is to examine the field's attention to common critiques regarding insufficient credible valuations of realistic changes to services; an unjustified preoccupation with monetary valuation; and too little social and policy research (e.g. questions of access to and demand for services). We found that very little of the ES literature includes valuation of biophysical change (2.4%), despite many biophysical studies of services (24%). An initially small but substantially rising number of papers address crucial policy (14%) and social dimensions, including access, demand and the social consequences of change (5.8%). As well, recent years have seen a significant increase in non‐monetary valuation (from 0% to 2.5%). Ecosystem service research has, we summarize, evolved in meaningful ways. But some of its goals remain unmet, despite the promise to improve environmental decisions, in part because of a continued pre‐occupation with numerical valuation often without appropriate biophysical grounding. Here we call for a next generation of research: Integrative biophysical‐social research that characterizes ES change, and is coupled with multi‐metric and qualitative valuation, and context‐appropriate decision‐making. A free Plain Language Summary can be found within the Supporting Information of this article. A free Plain Language Summary can be found within the Supporting Information of this article.
Resource-selection studies of passerine birds during the breeding season have mainly been limited to understanding those factors important to nesting. However, little is known about what resources are selected by juveniles that are no longer dependent on their parents. The postfledging period may be a critical part of the breeding season for independent juveniles because they must avoid predators and learn to forage on a changing resource base. We used radio-telemetry to study postfledging habitat use and resource selection of juvenile Swainson's Thrushes (Catharus ustulatus) in coastal California from 2000 to 2002. We generated population-level contours (50% and 95% fixed-kernel) to describe habitat use by independent juveniles, and we determined juvenile resource selection by comparing vegetation characteristics at sites used by juveniles versus random sites. Juvenile Swainson's Thrushes used mixed-hardwood forest and coastal scrub during the postfledging period as well as riparian vegetation used by nesting adults. The most parsimonious predictors of resource selection were fruit abundance variables, suggesting that postfledging habitat selection by the Swainson's Thrush is best explained by the optimal-foraging hypothesis. We suggest that juvenile thrushes can track food resources in a habitat mosaic and use vegetation types distinct from what is traditionally considered Swainson's Thrush breeding habitat. Selección de Recursos por Juveniles de Cathatus ustulatus Durante el Periodo de Emancipación Resumen. Los estudios de selección de recursos en aves paserinas durante la estación reproductiva, se han centrado principalmente en entender los factores importantes para la anidación. Sin embargo, se sabe poco acerca de los recursos seleccionados por los juveniles que ya no dependen de sus padres. Para los juveniles independientes, el periodo de emancipación podría ser una parte crítica de la estación reproductiva en la que deben evitar depredadores y aprender a forrajear sobre una base de recursos fluctuante. Por medio de telemetría, estudiamos el uso de hábitat y selección de recursos en juveniles del zorzal Catharus ustulatus en la costa de California entre los años 2000 y 2002. A nivel de población, generamos polígonos (de 50% y 95% de “kernel” fijo) para describir la utilización de hábitat por juveniles independientes, y para determinar la selección de recursos comparamos las características de la vegetación de los sitios utilizados con la de sitios aleatorios pareados. Los juveniles utilizaron bosques mixtos y matorral costero durante el periodo de emancipación, así como la vegetación riparia utilizada por adultos nidificantes. Las variables con las predicciones más parsimoniosas del uso de recursos fueron las relacionadas con la abundancia de frutos, lo que sugiere que la selección de hábitat de emancipación por C. ustulatus es explicada en mejores términos por la hipótesis de forrajeo óptimo. Sugerimos que los zorzales juveniles pueden encontrar recursos alimenticios en mosaicos de hábitat y usar tipos de vegetación distintos a los que tradicionalmente se consideran como hábitat de anidación.
Streambank stabilization reduces flooding to farmland and streambanks, improves water quality, and helps prevent sedimentation. In Illinois, about half of the grain produced is barged down the Illinois River. According to one study led by the Illinois State Water Survey, about 30 to 50 percent of the sediment load choking the Illinois River comes directly from eroding feeder streams. This finding has led to a concerted effort in the last 15 years, to develop and test five highly effective, reliable and low-maintenance streambank stabilization methods.
Revegetation can either provide buffer strips around existing habitat remnants to protect them from external impacts, corridors between them to increase connectivity, or additional habitat to increase the area of vegetation available, or can enhance degraded remnant areas. Revegetation will also help retain biodiversity indirectly if it helps stabliize an otherwise degrading agricultural landscape. Revegetation in agricultural areas is compared with minesite rehabilitation, where the redevelopment of functioning ecosystems and faunal habitat appears to be possible. The task is harder in the agricultural situation because the scale of modification is greater, soil changes are more difficult to redress, and recolonization by native species is less likely. -from Author
In recent years an increasing share of fishery management resources has been committed to alteration offish habitat with artificial stream structures. We evaluated rates and causes of physical impairment or failure for 161 fish habitat structures in 15 streams in southwest Oregon and southwest Washington, following a flood of a magnitude that recurs every 2–10 years. The incidence of functional impairment and outright failure varied widely among streams; the median failure rate was 18.5% and the median damage rate (impairment plus failure) was 60%. Modes of failure were diverse and bore no simple relationship to structure design. Damage was frequent in low-gradient stream segments and widespread in streams with signs of recent watershed disturbance, high sediment loads, and unstable channels. Comparison of estimated 5–10-year damage rates from 46 projects throughout western Oregon and southwest Washington showed high but variable rates (median, 14%; range, 0–100%) in regions where peak discharge at 10-year recurrence intervals has exceeded 1.0 m3·s–1·km–2. Results suggest that commonly prescribed structural modifications often are inappropriate and counterproductive in streams with high or elevated sediment loads, high peak flows, or highly erodible bank materials. Restoration of fourth-order and larger alluvial valley streams, which have the greatest potential for fish production in the Pacific Northwest, will require reestablishment of natural watershed and riparian processes over the long term.
In the active floodplain of Cottonwood Creek, an alluvial stream in California, conditions for the establishment and growth of plants are largely controlled by periodic flooding. Flooding creates and modifies erosional and depositional surfaces which are occupied by different species. To evaluate patterns of species occurrence and dominance, geomorphic surfaces on Cottonwood Creek's floodplain were defined on the basis of vertical and horizontal position relative to the stream (i.e., flood frequency), microtopography and particle size. Vegetation sampling on these surfaces provided data for classifying communities by relative cover of common riparian species. The results indicated zonation of communities dominated by different species in relation to flood-induced disturbance. Plant communities dominated by Salix hindsiana and annual grasses were found on surfaces frequently flooded and subject to severe scouring or deposition. Populus fremontii attained dominance or shared dominance with S. hindsiana on less frequently flooded surfaces and where particle size indicated less disturbance by erosion or deposition. Juglans hindsii, Quercus lobata or mixed stands of P. fremontii/J. hindsii/Q. lobata/S. hindsiana were dominant on infrequently disturbed surfaces.
Many fish habitats have been altered in Pacific Northwest streams and rivers over the past century by a variety of land use practices, including forestry, urbanization, agriculture, and chan- nelization. There are research and management needs for evaluation of the effectiveness of rehabilitation projects intended to enhance stream fish habitat recovery. The response of populations of juvenile coho salmon Oncorhynchus kisutch and steelhead 0. mykiss to addition of large woody debris (LWD) was tested in North Fork Porter Creek (NFPC), a small coastal tributary of the Chehalis River, Washington. The NFPC was divided into three 500-m study sections; two sections were altered with two approaches (engineered and logger's choice) to adding LWD, and the third was kept as a reference site. Immediately after LWD addition, the abundance of LWD pieces was 7.9 times greater than the pretreatment level in the engineered site and 2.7 times greater in the logger's choice site; abundance was unchanged in the reference site. Subsequent winter storms brought additional LWD into all three study sites. In the years that followed, the amount of pool surface area increased significantly in both the engineered and logger's choice sites, while it decreased slightly in the reference site. After LWD addition, winter populations of juvenile coho salmon increased significantly in the engineered and logger's choice sites, while they remained the same in the reference site. There were no significant differences in the coho salmon populations during spring and autumn within the reference, engineered, or logger's choice sites. The coho salmon smolt yield from the engineered and logger's choice sites also increased significantly after LWD addition, while it decreased slightly in the reference site. After LWD addition, the reference site and the engineered site both exhibited increases in age-0 steelhead populations; however, the population in the logger's choice site did not change. There was no difference in age-l steelhead abundance among sites, or before and after enhancement during any season. Winter populations of juvenile coho salmon and age-0 steelhead were related inversely to maximum and mean winter discharge.