Invasive Species and Compensatory Wetland Mitigation

Item

Title

Eng Invasive Species and Compensatory Wetland Mitigation

Date

2006

Creator

Eng Ehorn, Casey H.

Subject

Eng Environmental Studies

extracted text

INVASIVE SPECIES AND COMPENSATORY
WETLAND MITIGATION SUCCESS

by
Casey H. Ehorn

A Thesis: Essay of Distinction

Submitted in partial fulfillment

of the requirements for the degree

Master of Environmental Studies

The Evergreen State College

June 2006


This Thesis for the Master of Environmental Studies degree

by

Casey Howard Ehorn


has been approved for

The Evergreen State College

by


Amy Cook, Ph.D.

.


Date

ABSTRACT

Invasive species and compensatory wetland mitigation

Casey Howard Ehorn

Polygonum cuspidatum, Lythrum salicaria, and Phalaris arundinacea are invasive plant
species that pose significant threats to the legal and functional success of compensatory
mitigation sites because they have the ability to form dense monostands. Many
compensatory mitigation wetlands fail to meet permit requirements because they exceed
the 10% standard for aerial coverage of invasive species, but may still be providing
functional replacement as required under the No-Net-Loss policy.

Table of Contents
List of Figures

List of Tables
Acknowledgements
Chapter 1
Introduction
Instruments of Corps Regulation
Wetlands
Species Invasions in Wetlands
Chapter 2: Japanese Knotweed (Polygonum cuspidatum)
History
Physical Description
Reproduction
Seeds
Chapter 3: Purple Loosestrife (Lythrum salicaria)
History
Physical Description
Seeds
Effects of monostands on habitats
Chapter 4: Reed Canarygrass (Phalaris arundinacea)
History
Agricultural Use
Physical Description
Seeds
Nutrient Enrichment.
Invasive Characteristics
Changes in hydrologic regime
Chapter 5: Discussion
Performance of Compensatory Mitigation Wetlands
Ecological success of compensatory mitigation wetlands
Conclusion
Policy recommendations
References

v

vi

vii

1

1

2

5

8

10

10

11

12

13

16

16

17

19

20

23

23

24

25

26

27

28

29

31

31

34

37

40

41


IV

List of Figures
Figure 1: Evaluation of 40 wetland mitigation projects in south Florida
Figure 2: Occurrence and predominance of invasive plant species

.33
36

v

List of Tables
Table 1: Summary of Corps permit decisions, by fiscal year.
Table 2: Wetland impacts authorized by Corps permit and wetland compensatory
mitigation required

.4

6

VI

Acknowledgements
I would like to thank my fiance, Ashlee for her unwavering support. Thank you to Amy
Cook, my reader for her advice and guidance. My utmost gratitude goes to my parents
for their love and encouragement.

Vll

Chapter 1

Introduction
In the last 200 years over 50% of the original wetlands in the United States were
destroyed (Dahl 1991; Mitsch and Gosselink 2000). As the public's understanding of
wetland functions and values increased, a variety of policies and laws were passed to
protect wetlands (Mitsch and Gosselink 2000). In 1972 Congress passed the Federal
Water Pollution Control Act, later amended as the Clean Water Act, which regulates the
placement of dredge and fill materials in the Waters of the United States, including
wetlands under Section 404. Compensatory mitigation is a significant part of the
Section 404 permitting process (Kruczynski 1990), not because it is required under the
Clean Water Act, but because the issuance of a Section 404 permit triggers mitigation
requirements under the National Environmental Policy Act (Berry and Dennison 1993).

Many compensatory mitigation wetlands are not considered successful because
they fail the 10% aerial coverage standard for invasive species. Invasive plants reduce
biodiversity in wetlands (Wilcove, et al. 1998). Invasive plants that reproduce
vegetatively and form monocultures are the most threatening to native plant communities
(Pysek 1997; Kercher, et al. 2005). When species composition shifts in compensatory
wetland mitigation some of the functions expected to be replaced by the mitigation
wetlands may be lost, which is in direct conflict with the concept of "no-net-1oss" of
wetlands and wetland functions. The 10% aerial coverage standard may not be

appropriate for all invasive species and all compensatory mitigation sites because
reference wetlands may also exceed the 10% aerial coverage standard.

Instruments of Corps Regulation
The Department of the Army has been involved in regulating certain activities in
the nation's waters since 1890. Initially the mission of the U.S. Army Corps of Engineers
(Corps) Regulatory program was to protect and maintain the navigability of the nation's
waters. In the late 1960's, the Corps' regulatory saw a dramatic change with the addition
of a second focus as a product of several new laws and court decisions. Today, the
Corps' mission is still evolving as public needs and policy change, and as case law and
new statutory mandates add to the complex character of the program's authority.

The legislative genesis of the regulatory program are the Rivers and Harbors Acts
of 1890 (superseded) and 1899, which establish permit requirements to prevent
unauthorized obstruction or alteration of any navigable water of the United States (33
U.S.c. 401, et seq.). The most regularly exercised authority is contained in Section 10
(33 U.S.c. 403) which covers "construction, excavation, or deposition of materials in,
over, or under such waters, or any work which would affect the course, location,
condition, or capacity of those waters."

In 1972, amendments to the Federal Water Pollution Control Act added Section
404 authority (33 U.S.c. 1344) to the Corps' regulatory program. Under section 404 of
the Clean Water Act, the Chief of Engineers, acting for the Secretary of the Army, is
authorized to issue permits for the discharge of dredged or fill material into waters of the

2

United States, including navigable waters and wetlands, at specified disposal sites after
giving notice of the proposed discharge and opportunity for the public to comment at
hearings.

In the 1975 decision of Natural Resources Defense Council v. Riverside Bayview

Homes, Inc., included wetlands in the definition of waters of the United States as defined
by the Federal Water Pollution Control Act, prior to the decision the Corps only regulated
Section 404 dredge and fill activities in navigable waters. In 1977 the Federal Water
Pollution Control Act was amended and given the name Clean Water Act, and the Act's
latest amendments in 1987 have changed criminal and civil penalties and added an
administrative penalty provision.

The selection of placement sites for dredge or fill material is done in accordance
with Section 404(b)(1) guidelines developed by the U.S. Environmental Protection
Agency (40 CFR Part 230). Navigable waters of the United States are those waters that
are subject to the ebb and flow of the tide, and/or are presently used, or have been used in
the past, or may be susceptible to use to transport interstate or foreign commerce. In
waters affected by the tide, the landward limit of the navigable water of the United States
is the mean high water; in non-tidal waters the landward limit of navigable waters is the
ordinary high water. In non-tidal waters where adjacent wetlands are present, the Clean
Water Act extends jurisdiction to the limits of the adjacent wetlands, as defined by the
1987 Corps of Engineers Wetland Delineation Manual (1987 Manual).

3


The Corps issues two types of Department of the Army permits, Standard and
General permits, with two permits in each category. The basic vehicle for authorization
used by the Corps is the standard individual permit. Standard permits include standard
individual permits and letters of permission, which both take into account the public
interest when processing a permit decision. General permits are issued on a nationwide
and regional basis to authorize categories of activities that are substantially similar in
nature and cause only a minimal individual or cumulative adverse effects on the aquatic
environment. There are currently 44 nationwide permits issued, with several either
expired or revoked on a regional basis. A summary of Corps permit decisions is provided
in Table 1.

Table 1: Summary of Corps permit decisions, by fiscal year
Permit Type

1999
4,168
2,687
44,913

individual permit
letters of permission
nationwide permit
regional general
permit
38,595
denials
221
Totals
Source: Engineers 2006

2000
3,883
2,560
41,385
40,702
180
88,710

Fiscal Year
2001
4,159
3,066
37,088
38,759
171
.

2002
4,023
3,258
35,768

2003
4,035
3,040
35,317

38,125
128
81,302

43,486
299
86,177

4


Wetlands
Wetlands have been dewatered and filled for farming, development, mosquito
control, and numerous other activities throughout the nation's history (Toxicology 2001).
Wetlands provide numerous ecosystem functions that are also lost when wetlands are
filled. These include water quality improvement, flood storage, ground water recharge,
shoreline stabilization, and habitat functions . In 1988 at the National Wetland Policy
Forum the concept of "no-net-Ioss" was introduced by the Conservation Foundation, and
was subsequently endorsed by the administration of President George H. W. Bush, and
articulated in the 1990 Memorandum of Agreement (MOA) between the U.S .
Environmental Protection Agency and the Corps. The Memorandum of Agreement set
up a mitigation sequence that recognized that wetlands provided important ecosystem
functions that were important to the goals of the Clean Water Act, which are to "restore
and maintain the chemical, physical, and biological integrity of the Nation ' s waters"
(33 U.S.c. 1251(a)). The U.S. Supreme Court has found that the larger goal of the Clean
Water Act is the improvement of water quality, and that wetlands adjacent to navigable
waters "playa key role in protecting and enhancing water quality.. . [and] serve
significant natural biological functions" (States 2002). Working with the Memorandum
of Agreement as a framework, the Corps stated that the goal of compensatory mitigation
was to replace the affected aquatic resource functions that will be lost or impaired by the
project, or to maintain or improve the overall aquatic environment (Corps 2006).

5


Mitigation may include avoiding, minimizing, rectifying, reducing, or
compensating for resources that will be lost due to the construction of the proposed
project. The Corps regulations state that losses will be avoided to the extent practicable,
but impacts for many projects are unavoidable. 33 CFR 320.4(r) Provides general
guidance about mitigation that is required for the Corps regulatory program; that
mitigation will be directly related to the impacts of the proposed project, appropriate in
scope and degree of impacts, and that the mitigation can be reasonably enforced.
Mitigation is an important concern when evaluating Department of the Army permit
applications because mitigation can be required to ensure that the project is adequately
compensating for the impacted aquatic resources (40 CFR Part 230.

Table 2: Wetland impacts authorized by Corps permit and wetland compensatory
. d
rru 19a IOn reqUIre
Fiscal
Year

Wetland impacts
permitted (acres)

1999
2000
2001
2002
2003
Source: Engineers

21,556
18,900
24,089
24,651
21,413

Wetland compensatory
mitigation required
(acres)
46,433
44,757
43,832
57,821
43,550

After an application for a Department of the Army permit has been considered
and either found to be consistent with activities already authorized under general permits,
or not found to be contrary to the public interest and otherwise compliant with Corps
regulations, a permit is issued contingent on appropriate approved mitigation being
constructed if the impact threshold is over 0.1 acres.

6


The 1990 MOA stipulates that the Corps consider the functional values lost when
determining compensatory mitigation requirements for a section 404 permit, and states
that in-kind compensatory mitigation is generally preferable to out-of-kind mitigation.
Therefore, under the section 404 permit review process, the Corps must attempt to
achieve replacement of the impacted or lost wetland functions and values.

Ecological performance standards are used to establish that the approved
compensatory mitigation is developing in the desired aquatic habitat and providing the
expected ecological functions. To ensure the success of the approved compensatory
mitigation, District Engineers may impose administrative and adaptive management
requirements including: as-built surveys, performance bonds, real estate instruments for
protection of mitigation sites, and long term management funding. These administrative
requirements are intended to guarantee that the compensatory mitigation site is
constructed as approved by the District Engineer, and that the mitigation is maintained
and protected from future development. Adaptive management requirements may
include modifications to management and maintenance of compensatory mitigation sites
based on monitoring of ecological performance standards.

All compensatory mitigation is usually held to some type of performance
standards (Engineers 2006). These standards are normally based on aquatic community
structure and aquatic resource function as they relate to the criteria in the 1987 Manual
for wetland hydrology, soils, and vegetation.

7


Species Invasions in Wetlands

In some cases, permit compliance is determined by survival of specific plantings.
Plant species introductions from other eco-regions have put a strain on native plant
communities in North America, and invasive species in wetlands can pose definite
challenges for compensatory mitigation being able to meet aquatic community structure
performance standards (Kennedy, et al. 2002). Mitchell and Gopal found that "there is
some validity to the concept that disturbed ecosystems are the most susceptible to alien
invasions," (Mitchell and GopaI1991).

Wetland mitigation sites are very susceptible to species invasions because they
are typically devoid of vegetation, have multiple gaps within plant canopies, and have
eutrophic water supplies (Toxicology 2001). Hobbs and Huenneke found that the habitat
being colonized by an invasive species will be more invadable if there are gaps in the
canopy or minor soil disturbances available for seedling colonization, conditions which
are typical during and after construction at wetland mitigation sites (Hobbs and Huenneke
1992). Plants that invade compensatory mitigation wetlands are usually species with
high seedling production and germination rates, and have the ability to spread
vegetatively (Toxicology 2001).

The most frequently encountered invasive species in the Pacific Northwest is reed
canarygrass (Phalaris arundinacea), a species that is difficult to eliminate because it
spreads by both seeds and rhizomes, and creates monocultures that crowd out lower
growing native plants (Merigliano and Lesica 1998). There is also great potential in the

8

Pacific Northwest for purple loosestrife (Lythrum salicaria), a tall emergent hydrophyte,
to creep onto
of many wetlands in the

sites, and this species has already shifted the species composition
and is causing alarm among wildlife managers

(Stucky 1980; Balogh and Bookhout 1989). But the invasive species with the most
invasion potential for compensatory mitigation sites is Japanese knotweed (Polygonum
cuspidatum), which poses many hazards to successful stream and riparian restoration
projects, and which has already begun to appear in the Pacific Northwest, some in
patches as large as half an acre (SolI and Morgan Undated).

9

Chapter 2: Japanese Knotweed (Polygonum cuspidatum)

History

Polygonum cuspidatum was first described in 1777 by Houttuyn as Reynoutria
japonica and as Polygonum cuspidatum by Siebold in 1846 (Beerling, et al. 1994). It was

not until the early part of the 1900's that these were discovered to be the same plant
(Beerling, et al. 1994). The plant is referred to as Polygonum cuspidatum by Asian and
American authors and as Fallopia japonica by European authors.

Polygonum cuspidatum is native to eastern Asia. It was introduced to the United

in 1825 as an ornamental (Townsend 1997), and in the late 1800's to North
America as an ornamental and fodder plant, but rhizomes are reported to be toxic to some
animals (Patterson 1976; Conolly 1977; Beerling, et al. 1994). Today in Japan it is used
to hide garbage dumps and shield other unsightly areas as well as to stabilize seashore
areas vulnerable to wave erosion (Locandro 1978; Jennings and Fawcett 1980).
Polygonum cuspidatum is edible, newly emerged shoots can be used in salads, older

shoots can be stripped and prepared in a manner similar to rhubarb, and are said to have
an almond flavor (Kiple 2000; Doll and Doll 2002). Hu chang, the dried roots and stem
of P. cuspidatum, are used by traditional Chinese medicine practitioners to treat high
cholesterol and various other conditions (Huang 1999). Polygonum cuspidatum has also
been used as a laxative (Lewis and Elvin-Lewis 1977). Roots contain a phytochemical
called resveratrol, which his also found in red wine, that may shield against cancer and
cardiovascular disease by acting as an antioxidant, antimutagen, and anti-inflammatory

10

agent (Kimura and Okuda 2001). Aqueous extracts of P. cuspidatum have been found to
inhibit the formation of new blood vessels in vitro (Wang, et al. 2004). During World
War II, leaves of P. cuspidatum were used as a substitute for tobacco
(Beerling, et al. 1994).

In Japan, P. cuspidatum is a pioneer species in the primary succession of
volcanically disturbed slopes and is a colonizer in secondary succession of hilly or high
mountain ecosystems on sites with direct sun exposure (Kanai 1983; Hirose and Tateno
1984).

European authors have found evidence that clones may persist on a single site for
over 100 years (Pyek, et al. 2001). In North America, P. cuspidatum has been observed
from Nova Scotia to North Carolina, and is widely distributed in the Midwest and in the
coastal regions of Washington and Oregon, where it spreads along river banks, as well as
wetlands, along roads and fence lines, and in other disturbed areas (Muenscher 1955;
Pauly 1986).

Physical Description
Polygonum cuspidatum is an herbaceous perennial which can grow to a height of

10 feet. It is dioecious and reproduces by seed, but can also reproduce vegetatively by
large rhizomes, which can be 20 feet or longer. Well established plants develop a central
taproot. The hollow stems are simple and glabrous (non-haired) with thin membranous
sheaths that extend from the erect base.

The leaves of P. cuspidatum are broad and ovate, truncating to cuneate at the leaf base,
cuspidate at leaf apex, 5-15 cm long, 5-12 cm wide, with petioles 1-3 cm long. Flowers
are greenish-white, 2-3 mm long, compactly arranged in axillary panicles. Male flowers
have branched panicles on upright racemes with the distal end of the raceme in the
highest position; individual panicles usually point up; 8-10 stamens with longitudinally
dehiscing anthers. Female flowers are decumbent with the proximal end in the highest
position; 3 styles, fruiting 6-10 mm long calyx. Both male and female flowers possess
rudimentary organs of the other sex. Trigonous achenes are shiny black-brown and 3-4
mm long (Fernald 1950).

Reproduction
The primary mode of reproduction in the United States is through rhizomes which
can be 15-20 meters long and which are dispersed when fragments of rhizomes are
transported by water or more commonly when disturbed soil containing rhizomes is
placed as fill; shoots commonly emerge in April and growth rates can exceed 8 cm per
day (Locandro 1973; Conolly 1977; Locandro 1978). The capacity of P. cuspidatum
rhizomes to generate viable shoots is affected by the source of rhizome fragments, size
and depth in soil (Locandro 1973). Polygonum cuspidatum has been observed
regenerating from internodal tissue (Locandro, 1973), and rhizomes fragments buried 1
meter deep can produce viable plants and have been observed growing up from two
inches of impervious surface (Pridham and Bing 1975; Locandro 1978).

In Europe P. cuspidatum has been observed growing on soils with pH values
ranging from 3.0 to 8.0 (Grime, et al. 1988). In its native Japan, P. cuspidatum it has

12

been observed growing on sulphurous soils near volcanic fumaroles at pHs below 4.0
(Yoshioka 1974).

Seeds
Polygonum cuspidatum is both dioecious and gynodioecious, and has been

observed to be subdioecious in New England, with male and female flowers on separate
plants with males that sometimes set seed (Beerling, et al. 1994; Forman and Kesseli
2003). Polygonum cuspidatum flowers from August to September in North America

(Fernald 1950; Conolly 1977), and the main method of seed dispersal in North America
is wind (Maruta 1976). Hirose and Tateno found high levels of seed production but low
seedling survival in their 1984 study on Mt. Fuji, but noted that once a seedling had past
the three-leaf stage, the seedling typically survived (Hirose and Tateno 1984). Wild
plants in Asia are characteristically found in early successional environments, and first
and second year seedlings can be found growing next to adult plants (Maruta 1981;
Schnitzler and Muller 1998).

In Europe, seedling establishment in the wild has been noted, but several
documented cases have turned out to be hybrids between P. cuspidatum and
P. sachalinense (giant knotweed) (Bailey, et al. 1995). New genetic research has

concluded that virtually all non-hybrid P. cuspidatum (referred to as Fallopia japonica
var. japonica) in the United Kingdom are female, implying clonal growth (Hollingsworth

and Bailey 2000).

13


Outside of Asia, "seeds do not appear to be a significant mode of reproduction"
(Seiger 1995). Seiger found that 50% to 63 % of seeds collected in Washington, D.C.
germinated on a filter paper medium after a two year dormancy, while only 10% of seeds
with no dormancy period germinated (Seiger 1995). Seiger also noted that the seeds
collected appeared to be hybrids with Polygonum aubertii, and did not observed any
seedling establishment in the field (Seiger 1995).

Forman, et al. found high viability of P. cuspidatum seeds from Massachusetts
under various environmental treatments; seeds from the same parent were able to
germinate in the fall almost immediately after seed set, or enter a dormancy period and
germinate in spring (Forman and Kesseli 2003). Seeds in the Forman, et al. study
remained viable in winter conditions whether attached to the parental plant, covered with
soil, or exposed on the soil surface (Forman and Kesseli 2003). Forman, et al. did find
that seedlings that germinated underneath well-established stands of P. cuspidatum were
typically not able to survive because the canopy of the existing stands blocked most
sunlight (Forman and Kesseli 2003). However, the Forman, et al. seeds dispersed into
areas with open canopies did survive, and do not inevitably die at early stages as reported
by Locandro and Seiger; who focused on P. cuspidatum areas which were already heavily
infested and competition prevented any plant from growing, particularly seedlings
(Locandro 1973; Seiger 1995; Forman and Kesseli 2003).

Locandro found in his 5 year study of P. cuspidatum in New Jersey that female
plants often bore empty achenes, fertile males were rare, and plants that did germinate

14

seldom developed past the three-leaf stage and did not survive beyond mid-summer
(Locandro 1973). Forman, et al. reported drastically different findings in Massachusetts,
in both greenhouse and field observations female plants isolated from males had ovaries
that aborted with no remnant seed detected (Forman and Kesseli 2003). Additionally,
Foreman, et al. found at least one fertile male plant within pollinating distance of each
female plant, suggesting that isolation of female stands is not as dramatic as reported by
Locandro (Locandro 1973; Forman and Kesseli 2003).

When P. cuspidatum invades riparian sites, simplification of forest structure can
lead to decreases in small mammal habitat, and change nutrient cycling, prevent
recruitment of large woody debris, disrupt the aquatic food webs of salmoinds, block fish
passage, and simplify normally complex salmonid habitat (Potash 2001).

15


Chapter 3: Purple Loosestrife (Lythrum salicaria)

History

Purple loosestrife (Lythrum salicaria) is an emergent aquatic native to Europe and
Asia that was first described by Turner in 1548. Lythrum salicaria was transported to the
United States in the early 1800's in soil used as ballast for ships, livestock bedding, and
as a garden plant (Hulten 1971). Lythrum salicaria became established so quickly in the
costal eastern wetlands of North America that in the first edition of A Flora ofNorth

America, Torrey and Gray described L. salicaria as "probably native" (Voegtlin 1998).
Torrey noted in 1877 that L. salicaria was "well established on the Northern R[ail]
R[oad] of New Jersey, near Granton" (Torrey 1877). In 1879 L. salicaria was reported in
abundance along the high water mark of the Hudson River, and in meadows on adjacent
creeks (Rudkin 1879).

The colonization of L. salicaria into glacial marshes of the Midwest by 1900 is
correlated with development in these wetlands habitats; construction of eastern canals,
and marine traffic and trade extending into the Great Lakes region (Skinner, et al. 1994).
In the 1940s disturbed areas in which L. salicaria colonizes increased significantly as
construction commenced of the first series of interstate highways, and as more acreage
was irrigated under the Federal Reclamation Act (Kuusvouri 1960).

16


By the late 1940's, L. salicaria was established in marine areas in the Pacific
Northwest, and by 1985 Alaska and Montana were the only non-Southwestern states that
had not reported L. salicaria (Thompson, et al. 1987). In the more recent studies L.

salicaria was found to be invading new wetland sites by migrating down ditch and
culvert drainage systems (Wilcox 1989). There has also been interest by bee keepers in
using L. salicaria as a honey plant dating as far back as 1944, which could have
contributed to the species spread into the west (Thompson, et al. 1987). More recently L.

salicaria seeds have been found to be contaminating seed samples obtained from
commercial suppliers of wildlife cover and prairie restoration plants, undoubtedly
contributing to some current invasions (Thompson, et al. 1987).

Physical Description

Lythrum salicaria is a broad-leafed emergent aquatic perennial that can reach up
to 8 feet in height. Leaves are lanceolate, terminating to cordate, usually opposite, but
may also be alternate, or in whorls of three and four. An individual plant may have 30-50
angular annual stems that turn woody with age, rootstocks persist through winter
(Skinner, et al. 1994). Some uncertainty exists in the literature as to whether L. salicaria
rootstocks can send out rhizomes. Skinner, et al. said that L. salicaria may be
considered a clonal plant with the rootstock acting as genet, and annual shoots as ramets
(Skinner, et al. 1994). Ohwi described L. salicaria in Japan as "rhizomatous;" (Ohwi
1965) but this observation has been criticized as possibly referring to adventitious buds
sprouting from lateral roots (Thompson, et al. 1987). Pearsall described L. salicaria as

17

having "have tough rhizomes capable of penetrating the interstices of the hard
substratum" (Pearsall 1918), while other British authors of the time only mention
creeping rootstalks (Morse and Palmer 1925). In a more recent study the U.S. Fish and
Wildlife Service reported adventitious buds on the buried stems, but no evidence of
spread by rhizome activity after a thorough examination of plants throughout the United
States and Canada (Thompson, et al. 1987).

Lythrum salicaria plants bloom July to September and seed set begins by late
July. Spiked flowers are reddish-purple, six-petaled, showy, and have 5 sepals fused at
the base into a tube. The bisexual complete and perfect flowers are insect-pollinated;
self-pollination is achievable, but cross-pollination prevails (Thompson, et al. 1987).
Each flower has 8-10 stamens of 3 distinct lengths, and three distinct style lengths, one of
three pistil lengths and two different sets of stamen lengths in each flower. Shorter styles
may be hidden within the calyx (Stout 1925). The trimorphic nature of L. salicaria
flowers attracted the attention of Charles Darwin, who published two separate articles on
the subject. Darwin noted that the three forms of flowers coexisted in wild populations in
nearly equal frequencies (Darwin 1865). Kuusvouri found a high frequency of mid
length style morphs in crowded stands of L. salicaria in Finland where vegetative
reproduction was common (Kuusvouri 1960), but Halkka and Halkka while investigating
16 isolated populations of L. salicaria on small islands in the Gulf of Finland and
reported that findings similar to Darwin's; the three style morphs occurred in nearly 1:1 :1
frequencies (Halkka and Halkka 1974). The majority of the literature reported nearly
equal frequencies of all style morphs, supporting the assertion that sexual reproduction is

18

the driving factor in establishment and spread of L. salicaria (Shamsi and Whitehead
1974; Thompson, et al. 1987).

Seeds

The U.S. Fish and Wildlife Service reported the mean number of seeds produced
by a single plant to be 2,700,000, with 30 stems per plant, 1000 capsules per stem, and 90
seeds per capsule (Thompson, et al. 1987). The main mode of seedling dispersal is
floating seeds, which have been noted to sink upon contact with water, but rise after
germination (Ridley 1930), but seeds have also been observed being dispersed by wind
(Shamsi and Whitehead 1974), and by animals (Torrey 1931). Nilsson and Nilsson
suggest that L. salicaria seeds can be dispersed by wind over snow and ice (Nilsson and
Nilsson 1978). Thompson, et al. disputes this conclusion by noting that while seeds are
light enough to carried by wind, observed densities of seedlings fall off dramatically
within the first 10 m from the parent plant (Thompson, et al. 1987).

Shamsi and Whitehead found that 80% of seeds stored at 3°_4°C for three years
were able to germinate, while 90% of freshly collected seeds were able to germinate
(Shamsi and Whitehead 1974). Most L. salicaria seeds emerge by day 17, but low
reserve of stored energy by the seeds suggests that floating germinated seeds would not
survive beyond a few weeks (Shamsi and Whitehead 1974). Mitchell found that 75% of
L. salicaria seeds exposed to diffuse light successfully germinated, versus 6% of seeds

that were exposed to constant darkness (Mitchell 1926). Seeds need between
temperatures between 15 and 20°C to germinate, and can germinate successfully on

19

substrates with pHs between 4.0 and 9.1 (Shamsi and Whitehead 1974; Thompson, et al.
1987). Seeds or propagules that survive the winter must establish themselves in moist
soil in late spring or early summer; summer-germinated seedlings which do not produce
more than four pairs of leaves do not survive the following winter (Thompson, et al.
1987).

Seedlings of L. salicaria were found to be most affected by nitrogen deficiency,
as compared to phosphorus and potassium deficiency (Shamsi and Whitehead 1977).
Edwards, et al. noted that fecundity was similar among native L. salicaria populations in
Europe exposed to low and medium nutrient treatments, but was significantly higher at
both treatment levels amongst North American populations (Edwards, et al. 1998).

Effects of monostands on habitats

Many authors have also noted that invasions of L. salicaria often lead to changes
in arthropod, mammal and avian fauna that rely on native plants, shelter, or breeding and
nesting areas (Thompson, et al. 1987; Mal, et al. 1992; Kiviat 1996). This is because L.
salicaria has the ability to radically shift species composition within wetlands.

Fernald recognized the potential for invasion in 1940 when he penned that "

.

the formerly unique and endemic flora of the estuary is being rapidly obliterated by

.

the purple loosestrife ... without mercy for the insignificant endemics ...

II

(Fernald

1940). In 1978, Coddington and Field suggested that competition between L. salicaria

20

and Long's bulrush (Scirpus longii) may be partly to blame far the endangered status of

S. longii in Massachusetts (Coddington and Field 1978). Rawinski expressed similar
concern about a rare inland population of dwarf spike rush (Eleocharis parvula) in inland
New Yark (Rawinski 1982).

Rawinski and Malecki conducted a three year study and found a negative
correlation when comparing stem densities of L. salicaria and Typha (Rawinski and
Malecki 1984). In 1994 the presence of L. salicaria in flood, control and infertile
treatment plots caused an average 60% reduction in biomass of neighboring species
(Keddy, et al. 1994). A study of 12 Minnesota wetlands in 1995 showed that increased L.
salicaria biomass was associated with a decrease in Typha biomass (Emery and Perry

1995).

Weiher, et al. found that wetland microcosms inoculated with the seeds of 20
wetland plant species came to be dominated by L. salicaria after 5 years; most dicots had
vanished from the wetland microcosms and despite 24 different treatments Weiher and
colleagues concluded that initial planting conditions could not predict longer term trends
in species competition in the wetland microcosms (Weiher, et al. 1996). Another 1996
study by Twolan-Strutt and Keddy found that L. salicaria was less sensitive to
competition than Carex crinita by comparing biomass levels of roots and shoots in
competition treatments (Twolan-Strutt and Keddy 1996).

21

Mal, et al. found in a four year study with differing initial density treatments of L.

salicaria and Typha angustifolia that L. salicaria exceeded T. angustifolia in ramet
production after year one, exceeded stem proportion in all treatments after year four
(Mal, et al. 1997). Weihe and Neely had similar results when comparing L. salicaria and

Typha latifolia in varying light treatments; L. salicaria was able to produce more above­
and belowground biomass in all treatments (Weiher and Neely 1997). After herbicide
treatments (Gabor, et al. 1996) and cutting (Wilcox, et al. 1988) to remove L. salicaria ,
densities of graminoids have increased. Farnsworth and Ellis recognized that interspecies
competition was strong with L. salicaria; increasing L. salicaria biomass was linked with
declining biomass of other species, but questioned the frequency of true monostands of L.

salicaria (Farnsworth and Ellis 2001).

Brown studied the impact of L. salicaria on the native species L. alatum, and
found that pollinator visitation and subsequent seed set was lower in L. alatum with the
presence of L. salicaria (Brown 1999). Fickbohm and Zhu found that L. salicaria
transpired about twice as much water as Typha species, and concluded that monostands
of L. salicaria had the ability to cause changes in organic matter distribution, nitrogen
cycling, and water quality of freshwater wetlands (Fickbohm and Zhu 2006).

22


Chapter 4: Reed Canarygrass (Phalaris arundinacea)

History

Reed Canarygrass (Phalaris arundinacea) is a rhizomatic perennial that may be
native to North American costal areas, but that has spread under human influence
(Anderson 1961). It is reported to be native to Japan, Eurasia, North America, and South
Africa (Tsvelev 1983). It may be called 'canarygrass' either because the genus was first
described in the Canary Islands or because a close relative, P. canariensis is the source of
canary seed (Pojar and MacKinnon 1994). The first mention ofreed canarygrass was in
Hesselgren's 1749 thesis studying the preferred feeding species of livestock (Stannard
and Crowder 2001). European cultivation of reed canarygrass was documented in
England as early as 1824, and in Germany in 1850 (Schoth 1938).

Herbarium collections in the Pacific Northwest from the mid-1880's found this
species well represented, and the samples were frequently collected from remote
locations, indicating that P. arundinacea is native to North America (Merigliano and
Lesica 1998). Turner documented oral history that indicates Halq'emeylem and most
likely other Salish groups used stems for decorating baskets prior to European contact
(Turner 1992). But others speculate that freshwater wetland systems in Western
Washington prior to European contact were dominated by Thuja plicata, Picea sitchensis,
and Tsuga heterophylla, all of which would have created unfavorable understory
conditions for P. arundinacea (Antieau 1998), because species richness drops as intense

23

competition for available light increases during the stem exclusion stage of succession
(Houston, et al. 1996). However, many native American groups are noted to have
utilized burning of certain areas which would have maintained emergent communities
instead of forested wetland communities (Pyne 2001). Some native American groups
were even noted to burn wetlands for blueberry production (Adamson 1926).

Agricultural Use

In the 1830's P. arundinacea was used in livestock grazing trials on the Atlantic
Coast, most likely using local germplasm, and by the 1850's native reed canarygrass
stands were commonly used as grazing areas for livestock (Stannard and Crowder 2001).
As popularity of P. arundinacea as a grazing grass skyrocketed, European companies
began to export seed to North America, but most of the reed canarygrass currently
growing in the Pacific Northwest can be attributed to commercial seed production of
local germplasm in Coquille Valley of Oregon beginning in 1885 (Schoth 1938). The
species was noted to produce 30% more hay than similar grasses (Wilkins and Hughes
1932). Marten warned that P. arundinacea contains alkaloid compounds that at high
concentrations make it indigestible or toxic (Marten 1973).

Several authors have also noted that P. arundinacea agricultural stands do not
typically develop without multiple seedlings (Wasser 1982; Antieau 1998). This may be
due to the fact that P. arundinacea does not develop tillers until five to seven weeks after
germination (Comes, et al. 1981). Phalaris arundinacea's delay in sending up tillers may

24

limit it's initial completive ability against more rapidly tillering plants such as Agrostis
alba and Festuca rubra, but after tillering it gains a distinct competitive advantage (U.S.

Department of Agriculture 1996).

Use of P. arundinacea in the Pacific Northwest began in the late 1890's. Phalaris
arundinacea was used as a "breaking in" crop after logging (Wheeler 1950). In the late

1970's interest was sparked again in reed canarygrass as a wastewater management
species. Zeiders reported that "reed canarygrass is the most popular species for irrigation
with wastewater from municipal and industrial sources as a pollution control measure"
(Zeiders 1976). Phalaris arundinacea has also been utilized to stabilize shorelines and
prevent gully erosion (Baltensperger and Kalton 1958; Figiel, et al. 1995). More
recently, using P. arundinacea as a bio-fuel has been explored in Scandinavia (Katterer,
et al. 1998).

Physical Description

Phalaris arundinacea is a hollow stemmed, sod forming perennial, clums 0.7 to 2

meters tall, with scaly long rhizomes. Leaves are slightly hairy, flat, 5-15mm wide, 7 to
41cm long, with 4 to 10 mm membranous ligules, usually frayed and turned down.
Compact reddish panicle that changes to straw color as seeds mature; up to 25cm; up to 3
lanceolate spikelets per raceme, usually containing three florets, two of which infertile
and reduced. Slightly hairy glumes 4-5mm, and shiny lemma 4mm (infertile florets have
1mm lemmas).

25

Rhizomes arising from a single plant grow radially outwardly until a terminal bud
develops a shoot (Evans and Ely 1941). Comes, et al found that although P. arundinacea
develops thick rhizomes, they are relatively shallow-rooted; 88% of new shoots originate
from the upper Scm of soil, and 100% originate from upper 20cm of soil (Comes, et al.
1981). Whole plants dislodged during a disturbance event are able establish mono-stands
at new sites by re-rooting in disturbed soils (Hovin, et al. 1973). 74% of new shoots arise
from rhizomes, the remainder from auxiliary buds on basal nodes (Casler and Hovin
1980). The rhizomes of P. arundinacea are extremely tolerant of anoxiant conditions
(Brandle 1983).

Seeds

Seeds of P. arundinacea provide a means for long distance dispersal, exchange of
genetic information, and the impetus for multiple genotypes; multiple genotypes ensure
that at least some genotypes will thrive and multiply in harsh environments (Morrison
and Molofsky 1999). Seeds are naked, up to 3mm and germinate immediately after
ripening on long clumps, which ripen from top to bottom, allowing for a prolonged
period of dispersal. Seeds have no known dormancy requirements, and often dominate
seedbanks within wetlands (Apfelbaum and Sams 1987). In 97% of greenhouse grown P.
arundinacea seeds germinated immediately after harvest, while seeds stored in moist

sand germinated after a year of fluctuating temperatures (Comes, et al. 1981).

26

Seedlings are initially very sensitive to interspecies competition, and frequently
sprout in ephemeral ponds in spring (Morrison and Molofsky 1998). The ponded water
creates anaerobic conditions that force the sprouts to rely on carbohydrate reserves stored
in rhizomes (Hovin, et al. 1973). If water persists on the developing stand for extended
periods of time, the reducing environment will deprive the roots oxygen, killing the stand,
and in some cases removing oxygen from the roots (Stannard and Crowder 2001).
Individual leaves of P. arundinacea grow from nodes along the clum, and become
disadvantaged as the plant grows taller shielding lower leaves from light (Stannard and
Crowder 2001). Large mono-stands are capable of producing up to 9 tons/acre of
biomass, but this type of growth requires a tremendous amount of nutrients (Stannard and
Crowder 2001).

Nutrient Enrichment

Nutrient influx into wetlands can be associated with either agricultural or
residential runoff (Kercher, et al. 2005). Increases in runoff events can cause standing
water in wetlands, causing a decline in species that are not flood tolerant, making much
more light available for P. arundinacea (Kercher, et al. 2005). As the standing water is
released from the wetland, many of the nutrients and sediments that the runoff contained
increase growth in P. arundinacea. Kercher, et al. found that wetland mesocosms
subjected to flooding, sedimentation, and nutrification became monostands of P.
arundinacea after two growing seasons, with up to 50% of native species dying in the

first 6 weeks due to prolonged flooding (Kercher, et al. 2005).

27

Phalaris arundinacea has shown significantly increased growth associated with
nutrient enrichment of soil or water supply (Stannard and Crowder 2001). Ho found
increased stem density and increased nitrogen and phosphorus levels in P. arundinacea
that had been subjected to nutrient enrichment (Ho 1979). Both Green and Galatowitsch,
and Maurer and Zedler showed an increase in biomass with high nutrient treatments, and
Wetzel and van der Valk found a 73% increase in biomass of high nutrient treatments
over low nutrient treatments of P. arundinacea (Wetzel and van der Valk 1998; Green
and Galatowitsch 2001; Maurer and Zedler 2002). Maurer and Zedler did not find a
significant relationship between nutrient treatments and emergence or survival of P.

arundinacea communities, but did find that young ramets with nutrient treatments readily
invaded shaded areas drawing nutrients from parental clones (Maurer and Zedler 2002)..

Invasive Characteristics

Wetlands that experience invasion by P. arundinacea often have drastic declines
in native species within several years Spuhler found that wetlands with P. arundinacea
had 25-33% less species than neighboring sedge meadows, and at two sites P.

arundinacea had formed monotype stands (Spuhler 1994). Likewise, Werner showed
that wetlands invaded by P. arundinacea had 9 to 11 less species per square meter than
nearby wet prairie communities. Native herbaceous species that begin growing in late
spring can be dramatically affected by large monoculture stands of P. arundinacea,
which deprives them of light (Stannard and Crowder 2001). On the other hand
Lindig-Cisneros and Zedler found that sites that support dense native plant canopies

28

because of ideal site conditions can inhibit P. arundinacea invasion from seeds, even
when the native plant communities are not diverse(Lindig-Cisneros and Zedler 2002).
Lindig-Cisneros and Zedler's results also showed that among native plant canopies that
experienced a disturbance, canopies with higher diversity (more species) had greater
resistance to invasion by P. arundinacea seeds than native monotype canopies (Lindig­
Cisneros and Zedler 2002). Due to low tissue density P. arundinacea stems are twice as
high as similar grasses when grown alone, even thought biomass allocation is similar; but
when grown with native grasses P. arundinacea is able to change morphology and
increase its total shoot length to biomass ratio by 50% (Miller and Zedler 2003).

Changes in hydrologic regime

The ability of P. arundinacea to invade sites that have experienced a change in
hydrologic regime has been well documented. Apfelbaum, et al. considered P.

arundinacea a forceful invader in disturbed habitats where the substrate was favorable
(Apfelbaum and Sams 1987). Odland conducted vegetation transects on a reservoir that
was subject to a permanent drawdown and observed that P. arundinacea gradually
invaded not only the newly exposed substrate, but also the wetlands previously adjacent
to the reservoir (Odland 2002). Barnes reported an expansion of P. arundinacea on small
river islands following lower summer flows that exposed more river substrate (Barnes
1999). Good, et al. and Lech found increased germination and growth of P. arundinacea
after a hydrologic drawdown (Good, et al. 1978; Lech 1996). Barnes noted that riparian
areas may be particularly susceptible to invasion for two different reasons: because

29

sedimentation caused by flooding regularly makes available new sites for P. arundinacea
to establish; and because human activities along rivers can change the hydrologic regime
and alter rates of erosion and sedimentation, and disturb existing vegetation (Barnes
1999). Similarly, Comes, et al. found that P. arundinacea is a ready invader at sites
disturbed with chemical or mechanical control treatments which open up canopies
(Comes, et al. 1981).

Phalaris arundinacea has been found to be productive under varying moisture
levels (Morrison and Molofsky 1998), including flooding (Figiel, et al. 1995) and drought
(Sheaffer, et al. 1992). Miller and Zedler found that in flood conditions P. arundinacea
allocated more biomass above ground (Miller and Zedler 2003). Rubio and Lavado
suggest that this may be a mechanism to decrease biomass and oxygen demand of the
root systems, or to increase the ratio of root length to root biomass to aid in nutrient
uptake ability (Rubio and Lavado 1999). In droughty conditions P. arundinacea reduces
leaf surface area, heavily controls stomatal transpiration, and produces smaller cells with
thicker cell walls which retain more water than larger thin walled cells (Frank, et al.
1996).

30

Chapter 5: Discussion

Performance of Compensatory Mitigation Wetlands

Roberts summarized the process of permitting wetland impacts and requiring
compensatory mitigation by stating that "wetland trading is a loser's game"(Roberts
1993). Indeed, by simply evaluating compliance under the no-net-Ioss concept, wetland
losses continue to happen because of the failure of compensatory wetland mitigation sites
to be constructed. The U.S. EPA estimates that the United States is still loosing 70,000­
90,000 acres per year, which does not take into account the losses from compensatory
mitigation wetlands that are poorly designed or managed and therefore have reduced
functional value that does not adequately compensate for the aquatic resources that were
impacted (Lee and Chapman 2001).

Some types of herbaceous wetlands, such as freshwater emergent marshes and
wet meadows, have been successfully restored or created for compensatory mitigation
(Lindau and Rossner 1981; Niswander and Mitsch 1995; Wilson and Mitsch 1996;
Brown and Veneman 1998). Wet prairies and sedge meadows have met with limited
success (Galatowitsch and Valk 1996; Ashworth 1997). Shrub swamps and forested
wetlands are the most difficult to create or restore for compensatory mitigation because
of the time required to establish mature woody plants (Niswander and Mitsch 1995;
Brown and Veneman 1998; King 2000).

Most studies suggest that there is much room for improvement in the construction
of compensatory mitigation wetlands. Maguire found that 50% of mitigation sites in
Virginia were considered successful using area, vegetative cover, and achievement of
permit conditions to calculate mitigation success (Maguire 1985). Maguire noted that
many mitigation efforts were not considered to be successful because they had not even
been built (Maguire 1985). The U.S . Environmental Protection Agency found similar
results (Reimold and Cobler 1985). Glubiak, et al. and Quammen both suggested the
need for better management of compensatory mitigation wetlands (Glubiak, et al. 1986;
Quammen 1986).

There is currently a disagreement among researchers as to the success of recent
mitigation efforts. Harvey and Josselyn believed that compensatory mitigation was
working in the mid 1980' s, while Race suggested that many mitigation projects were frail
and could easily fail (Harvey and Josselyn 1986; Race 1986). Kusler and Groman
questioned the granting of permits to projects that were not water dependent, and when
alternative sites were available, and Golet took a firmer stance and asserted that damage
to wetlands should not be permitted unless there is absolutely no alternative (Golet 1986;
Kusler and Groman 1986).

Recent research has shown that many mitigation sites are not constructed. Erwin
(1991) found that in southern Florida only half of the 430 ha of wetlands required as
mitigation had been constructed, and 60% (24 of 40) of projects were found to be
incomplete or failures (Figure 1). Kentula et al. found similar results for mitigation sites

32

Ecological success of compensatory mitigation wetlands

Many researchers suggest that compensatory mitigation wetlands do not
adequately replace the structure and functions of the natural wetlands that are lost
(Streever 1999). Confer and Niering compared five created compensatory mitigation
wetlands with five natural ones in the same area and found more open water area in the
created wetlands, and also found that the source of hydrology for the created wetlands
was dependent on highway runoff (Confer and Niering 1992). Zedler and Malakoff
declared a 12 ha salt marsh mitigation site in southern California a failure after ten years
of monitoring because it did not provide habitat for the endangered light-footed clapper
rail (Rallus longirostris levipes) (Zedler 1996; Malakoff 1998). Wilson and Mitsch
concluded that while 4 of the 5 compensatory mitigation wetlands they studied were
successful ecologically, they were not always considered successful legally (in
compliance with permit conditions) (Wilson and Mitsch 1996). While Svengsouk and
Mitsch found that a 6 ha compensatory mitigation wetland in Ohio was successful, a
similar site in lllinois was not (Mitsch and Flanagan 1997; Svengsouk and Mitsch 1997).

Invasive species pose numerous challenges for compensatory mitigation. Many
compensatory mitigation wetland sites are determined to be compliant once the
vegetative community becomes established (Council 2001). While there are many things
that could be used to monitor success of mitigation wetlands, vegetation is considered the
easiest indicator of progress to observe (Mitsch and Gosselink 2000). In most cases, the

34

wetland indicators of hydrology and soils are rarely used to evaluate compliance with
permit conditions.

In 1998 Gallihugh studied the success of wetland mitigation sites in the Chicago
region by evaluating 61 permits issued from 1990 to 1994 with a combined impact of
288.7 acres (Gallihugh 1998). The total mitigation proposed under the 61 permits was
354.9 acres, but 72.5 acres of mitigation was never established, a net loss of 6.3 acres. Of
projects with constructed mitigation, 54 of 61 permits had special conditions related to
mitigation, but only 2 of these permits were deemed to be in full compliance with all
special conditions. Furthermore, 128 mitigation sites were required by the issuance of
the 61 permits; 22 of those sites were established with correct plant communities as
proposed, 28 mitigation sites had established wetlands, but with the wrong plant
communities, and the remaining 78 sites either had excessive open water or insufficient
hydrology to support the correct plant communities. The Chicago study also found that
64% of wetland mitigation sites had less than 20% of the plants that had initially been
planted actually establish. The overall success rate for a given native plant to establish in
an area where it was planted was 12.4%; only 10% of native plants had a success rate
above 50%, and 68% of native plants had success rates of 0%. As illustrated in Figure 2,
Lythrum salicaria was found at 48 of the 128 mitigation sites, and had a total aerial
coverage of more than 20% at 6 sites, while Phalaris arundinacea was found at 81 sites,
and had a total aerial coverage of more than 20% at 12 sites.

35


One example of a singular criterion used to judge compensatory wetland
mitigation success is the Floristic Quality Assessment developed by Swink and Wilhelm
for wetlands in the Chicago region and used by the Chicago District, U.S. Army Corps of
Engineers (Swink and Wilhelm 1994). The Floristic Quality Assessment essentially
characterizes a compensatory mitigation site solely on the vegetative community present
(Swink and Wilhelm 1994). The basic postulation underlying the Floristic Quality
Assessment approach is that specific wetland vegetation variables can be used to indicate
the functional success of compensatory mitigation, because if a vegetative community
displays a diverse pre-European condition, then the "physical, biological, and
biochemical functions that support the vegetation must be present" (Council 2001).
However, low plant diversity is not always characteristic of substandard hydro-geological
and geo-chemical conditions in wetlands, and higher plant diversity is not automatically a
de facto indicator of wetland functions (Council 1995).

Conclusion

The different types of floristic assemblages required for successful compensatory
mitigation often require widespread plantings and persistent management to maintain the
species composition desired (Council 2001). Polygonum cuspidatum, Lythrum salicaria,
and Phalaris arundinacea all pose significant threats to the legal success of
compensatory mitigation sites, because their aggressive growth often exceeds aerial
coverage standards that are made part of permit requirements. Often eradicating one of
these species from a site will cause disturbances which could allow new colonization by

37

yet another invasive species (Ecology 2004). Invasive species also present significant
challenges to the functional success of compensatory mitigation by changing species
composition, which can alter the functions that mitigation site was designed to replace.

The Corps use of the 10% aerial coverage standard for invasive species was an
attempt to implement a reasonable standard that could be easily measured, but the
scientific literature does not support the use of the 10% standard (Tong 2006).
Maintaining an aerial coverage of less than 10% for invasive species for the duration of
the monitoring period of a compensatory mitigation wetland does not ensure that a
mitigation site will not be invaded after the monitoring period ends, or that invasive
species already present take over a site in the future.

While the intent of the 10% standard for invasive species aerial coverage was to
prevent the establishment of monostands of invasive species from out competing native
species, thereby compromising and degrading wetland functions, many sites failed to
comply with the 10% standard and therefore permit conditions because of a high
occurrence of P. arundinacea (Celedonia 2002; Ecology 2004; Ecology, et al. 2006).
Because P. arundinacea does provide water quality, food web support, and habitat
functions (Terzi 2006), coverage standards for P. arundinacea should be set as to not
exceed aerial coverage at the impact site, or to not exceed the aerial coverage of nearby
wetlands, which ever is lower. Implementing this new standard will not result in a net­
loss of wetland functions, and will allow wetland mitigation to function in a manner that
is similar to nearby natural wetlands.

Polygonum cuspidatum and L. salicaria both form dense stands that are spreading
aggressively in many wetlands, displacing many native species(Mitsch and Gosselink
2000; Doll and Doll 2002). Polygonum cuspidatum also prevents woody species from
establishing on stream banks which disrupts aquatic food webs for salmonids (Potash
2001). For P. cuspidatum a zero tolerance policy should be adopted because of the
aggressiveness of recent invasions into wetlands, while the 10% aerial coverage standard
may be appropriate for L. salicaria because while it does invade wetland habitats, the
existence of true monostands of L. salicaria that exclude native species is questionable
(Keddy, et al. 1994; Farnsworth and Ellis 2001).

Polygonum cuspidatum, Lythrum salicaria, and Phalaris arundinacea, are
invasive species that form monostands which can undermine both the legal and
ecological success of compensatory mitigation wetlands by failing to replace the aquatic
resource functions that were impacted by the issuance of a Department of the Army
permit. Compensatory mitigation sites that exceed the 10% aerial coverage standard for
invasive species are not considered legally successful by the Corps because of permit
conditions specifying specific allowable invasive species coverage ratios. Compensatory
mitigation wetlands may also not be considered ecologically successful if the mitigation
site fails to provide the same wetland functions and values as the impacted site. A
diverse assemblage of native plants is always preferable at mitigation sites to ensure the
legal and functional ecologic success of compensatory mitigation wetlands.

39


Policy recommendations

• Adopt a zero tolerance policy for Polygonum cuspidatum.
• Make case by case determinations for aerial coverage standards of
Lythrum salicaria based on percent cover of natural wetlands close to the
mitigation site (not to exceed 10%).
• Set Phalaris arundinacea cover standards to match the impact site.
• Design aerial coverage standards for all compensatory mitigation wetlands on a
case by case basis and take into account the impact site, the mitigation site, and
the functions lost to be replaced by mitigation wetlands.

40


References
Adamson, T. (1926). Unarranged Sources of Chehalis Ethnology. Melville Jacobs
Collection, University of Washington Archives. Seattle, Washington: box 77,
parts I and II.
Anderson, D. E. (1961). "Taxonomy and distribution of the genus Phalaris." Iowa State
College Journal of Science 36: 1-96.
Antieau, C. J. (1998). Biology and management of Reed Canarygrass, and implications
for ecological restoration. W. S. D. o. Transportation.
Apfelbaum, S. 1. and C. E. Sams (1987). "Ecology and control of reed canarygrass."
Natural Areas Journal 7: 69-74.
Ashworth, S. M. (1997). "Comparison between restored and reference sedge meadow
wetlands in south-central Wisconsin." Wetlands 17(4): 518-527.
Atkinson, R. B., J. E. Perry, et al. (1993). "Use of created wetland delineation and
weighted averages as a component of assessment." Wetlands 13: 185-193.
Bailey, J. P., L. E. Child, et al. (1995). Assessment of the genetic variation and spread of
British populations of Fallopia japonica and its hybrid Fallopia x bohemica. Plant
invasions: general aspects and special problems. P. Pysek, K. Prach, M. Rejmanek
and M. Wade. Amsterdam, Netherlands, SPB Academic Publishing.
Balogh, G. R. and T. A. Bookhout (1989). "Purple loosestrife (Lythrum salicaria) in
Ohio's Lake Erie marshes." Ohio Journal of Science 89: 62-64.
Baltensperger, A. A. and R. R. Kalton (1958). "Varibility in reed canarygrass, Phalaris
arundinacea L. 1. Agronomic characteristics." Agronomy Journal 7: 659-663.
Barnes, W. J. (1999) . "The rapid growth of a population of reed canarygrass (Phalaris
arundinacea L.) and its impact on some riverbottom herbs." Journal of the Torrey
Botanical Society 126: 133-138.
Beerling, D. J., J. P. Bailey, et al. (1994). "Fallopiajaponica (Houtt.) Ronse Decraene."
Journal of Ecology 82(183): 959-979.
Berry, J. F. and M. S. Dennison (1993). Wetland Mitigation. Wetlands: Guide to science,
law and technology. M. S. Dennison and J. F. Berry. Park Ridge, New Jersey,
Noyes Publications: 278-303.

41


Brandle, R. (1983). "Evolution der garungskapazitat in den flut- und anoxiatoleranten
rhizomen von Phalaris arundinacea , Phragmites communis, Schoenoplectus
lacustris, und Typha latifolia." Boticancia Helvetica 93: 39-45.
Brown, B. J. (1999). The impact of an invasive species (Lythrum salicaria) on pollination
and reproduction of a native species (L. alatum). Kent, Ohio, Kent State
University. Ph.D. Thesis.
Brown, S. and P. Veneman (1998). Compensatory Wetland Mitigation in Massachusetts.
Research Bulletin 746. Amherst, Massachusetts, Massachusetts Agriculture
Experiment Station, University of Massachusetts.
Casler, M. D. and A. W. Hovin (1980). "Genetics of vegetative stand establishment
characteristics in reed canarygrass clones." Crop Science 20: 511-515.
Celedonia, M. T. (2002). Establishing appropriate benchmarks for site development by
documenting seccessional characteristics, Phase 2. R. a. S. D. U. Washington
State Department of Transportation, Washington State Department of
Transportation, Olympia, Washington.
Coddington, J. and K. G. Field (1978). Rare and endangered vascular plant species in
Massachusetts. Cambridge, Massachusetts, Committee for Rare and Endangered
Species of the New England Botanical Club: 62.
Comes, R. D., L. Y. Marquis, et al. (1981). "Response of seedlings of three perennial
grasses to dalapon, amitrol, and glyphosate." Weed Science 29: 619-621.
Confer, S. R. and W . A. Niering (1992). "Comparison of created and natural freshwater
emergent wetlands in Connecticut (USA)." Wetlands Ecology and Management
2: 143-156.
Conolly, A. P. (1977). "The distribution and history in the British Isles of some alien
species of Polygonum and Reynoutria." Watsonia 11: 291-311.
Council, N. R. (1995). Wetlands: Characteristics and Boundaries. Washington, DC,
National Academy Press.
Council, N. R. (2001). Compensating for Wetland Losses Under the Clean Water Act
Washington, D.C., National Academies Press.
Dahl, T. E. (1991). Wetland losses in the United States 1780's to 1980's. U. S. F. a. W. S.
U.S. Department of the Interior, Washington, D.C.
Darwin, C. (1865). "On the sexual relations of the three forms of Lythrum Salicaria."
Jorn. Linn. Soc. Bot. 8: 169-196.

42

Doll, J. and J. Doll (2002). "Japanese Knotweed (Polygonum cuspidatum)." Weed
Science.
Ecology, W. S. D. o. (2004). Reed canarygrass cover success standards for WSDOT
wetland mitigation sites in Washington, Washington State Department of
Ecology.
Ecology, W. S. D.o., U. S. A. C. o. E. S. District, et al. (2006). Wetland Mitigation in
Washington State - Part 1: Agency Policies and Guidance (Version 1),
Washington State Department of Ecology, Olympia, Washington.
Edwards, K. R., M. S. Adams, et al. (1998). "Differences between European native and
American invasive populations of Lythrum salicaria." Journal of Vegetation
Science 9: 267-280.
Emery, S. L. and J. A. Perry (1995). "Aboveground biomass and phosphorus
concentrations of Lythrum salicaria (Purple Loosestrife) and Typha spp. (Cattail)
in 12 Minnesota wetlands." American Midland Naturalist 134: 394-399.
Engineers, U. S. A. C. o. (2006). Draft Environmental Assessment, Finding of No
Significant Impact, and Regulatory Analysis for Proposed Compensatory
Mitigation Regulation. O. a. R. C. o. P. Directorate of Civil Works. Washington,
D.C.
Erwin, K. (1991). An evaluation of wetland mitigation in the South Florida Water
Management District. Volume 1. West Palm Beach, Florida, South Florida Water
Management District.
Evans, M. W. and J. E. Ely (1941). "Growth habits ofreed canarygrass." Journal of the
American Society of Agronomy 33: 1017-1027.
Farnsworth, E. J. and D. R. Ellis (2001). "Is purple loosestrife (Lythrum salicaria) an
invasive threat to freshwater wetlands? Conflicting evidence from several
ecological metrics." Wetlands 21(2): 199-209.
Fernald, M. L. (1940). "The problem of conserving rare native plants." Smithsonian
Institution Annual Report: 375-391.
Fernald, M. L. (1950). Gray's Manual of Botany 8th ed. Portland, Or., Dioscorides Press.
Fickbohm, S. S. and W. X. Zhu (2006). "Exotic purple loosestrife invasion of native
cattail freshwater wetlands: Effects on organic matter distribution and soil
nitrogen cycling." Applied Soil Ecology 32( 1): 123-131.

43


Figiel, C. R., B. Collins, et al. (1995). "Variation in survival and biomass of two wetland
grasses at different nutrient and water levels over a six week period." Bulletin of
the Torrey Botanical Club 122: 24-29.
Forman, J. and R. V. Kesseli (2003). "Sexual reproduction in the invasive species
Fallopiajaponica (Polygonaceae)." American Journal of Botany 90: 586-592.
Frank, A. B., S. Bittman, et al. (1996). "Water relations of cool-season grasses'."
Agronomy Monograph 34: 128-166.
Gabor, T. S., T. Haagsma, et al. (1996). "Wetland plant responses to varying degrees of
purple loosestrife removal in southeastern Ontario, Canada." Wetlands 16: 95-98.
Galatowitsch, S. M. and A. G. v. d. Valk (1996). "Characteristics of recently of recently
restored wetlands in the prairie pothole region." Wetlands 16(1): 75-83.
Gallihugh, J. L. (1998). Wetland mitigation and 404 permit compliance study. U. S. F. a.
W. S. U.S. Department of the Interior, Chicago, Illinois, U.S. Government
Printing Office. Volume I: Report and Appendices.
Glubiak, P. G., R. H. Nowka, et al. (1986). "Federal and state management of inland
wetlands: are states ready to assume control?" Environmental Management 10:
145-156.
Golet, F. C. (1986). "Critical issies in wetland mitigation: a scientific perspective."
National Wetlands Newsletter 8: 3-6.
Good, R. E., D. F. Whigham, et al. (1978). Freshwater wetlands. Ecological process and
management potential. New York, Academic Press.
Green, E. K. and S. M. Galatowitsch (2001). "Differences in wetland plant community
establishment with additions of nitrate-N and invasive species (Phalaris
arundinacea and Typha x glauca)." Canadian Journal of Botany 79: 170-178.
Grime, J. P., J. G. Hodgson, et al. (1988). Comparative Plant Ecology. London, Unwit
Hyman.
Halkka and L. Halkka (1974). "Polymorphic Balance in Small Island Populations of
Lythrum salicaria." Annales Botanici Fennici 11: 267-270.
Harvey, H. T. and M. N. Josselyn (1986). "Wetlands restoration and mitigation policies:
comment." Environmental Management 10: 567-569.
Hirose, T. and M. Tateno (1984). "Soil nitrogen patterns induced by colonization of
Polygonum cuspidatum on Mt. Fuji." Oecologia 61(2): 218-223.

44

Ho, Y. B. (1979). "Growth, clorophyll and mineral nutrient studies on Phalaris
arundinacea L. in three Scottish lochs." Hydrobiologia 63: 33-43.
Hobbs, R. J. and L. F. Huenneke (1992). "Disturbance, diversity and invasion:
implications for conservation." Conservation Biology 6: 324-337.
Holland, C. C. and M . E. Kentula (1992). "Impacts of Section 404 permits requiring
compensatory mitigation on wetlands in California (USA)." Wetlands Ecology
and Management 2: 26-29.
Hollingsworth, M. L. and J. P. Bailey (2000). "Evidence for massive clonal growth in the
invasive weed Fallopia japonica (Japanese Knotweed)." Botanical Journal of the
Linnean Society 133: 463-472.
Houston, D. B., E. S. Schreiner, et al. (1996). Effects of Fire Suppression on Ecosystems
and Diversity. B. R. D. U.S. Geological Survey, Forest and Rangeland Ecosystem
Science Center.
Hovin, A W., B. E. Beck, et al. (1973). "Propagation of reed canarygrass (Phalaris
arundinacea) from clum segments." Crop Science 13: 747-749.
Huang, K. C. (1999). The pharmacology of Chinese herbs, 2nd edition. Boca Raton, CRC
Press.
Hulten, E. (1971). The Circumpolar Plants. Stockholm, Almqvist & Wiksell.
Jennings, V. M. and R. S. Fawcett (1980). "Japanese Polygonum." Iowa State Univ.
Coop. Exten. Ser. Pm. 762. 2 p.
Kanai, H . (1983) . "Study on the distribution patterns of Japanese plants (5):
Accumulation of phytogeographical data of popular plants, points and measures."
Journ. Jap. Bot. 59: 257-269.
Katterer, T., O. Andren, et al. (1998). "Growth of and nitrogen dynamics of reed
canarygrass (Phalaris arundinacea L.) subjected to daily fertilization and irrigation
in the field." Field Crops Research 55: 153-164.
Keddy, P. A, L. Twolan-Strutt, et al. (1994). "Competitive effect and response rankings
in 20 wetland plants: are they consistent across three environments?" Journal of
Ecology 82: 635-643.
Kennedy, T. A, S. Naeem, et al. (2002) . "Biodiversity as a barrier to ecological
invasion." Nature 417: 636-638.

45


Kentula, M. E., J. C. Sifneos, et al. (1992). "Trends and patterns in Section 404
permitting requiring compensatory mitigation in Oregon and Washington, USA."
Environmental Management 16: 109-119.
Kercher, S., A. Herr-Turoff, et al. (2005) . What accelerates reed canarygrass invasions?
Madison, Wisconsin, University of Wisconsin: 1-4.
Kimura, Y. and H. Okuda (2001). "Resveratrol isolated from Polygonum cuspidatum root
prevents tumor growth and metastasis to lung and tumor-induced
neovascularization in Lewis lung carcinoma-bearing mice." Journal of Nutrition
131(6): 1844-1849.
King, S. L. (2000). "Restoring bottomland hardwood forests within a complex system."
National Wetlands Newsletter 22(1): 7-12.
Kiple, K. F. (2000). The Cambridge World History of Food. K. C. Ornelas. Cambridge,
Cambridge University Press: 1797.
Kiviat, E. (1996). "Tangled locks: the purple loosestrife invasion and biodiversity."
Annandale 5: 34-39.
Kruczynski, W. L. (1990). Mitigation and the section 404 program: A perspective.
Wetland creation and restoration: the status of the science. J. A. Kusler and M. E.
Kentula. Washington, D.C., Island Press.
Kusler, J. and H. Groman (1986). "Mitigation: an introduction." National Wetlands
Newsletter 8: 2-3.
Kuusvouri (1960). University of Helsinki, Finland. M.S. Thesis. in Thompson, Daniel Q.,
Ronald L. Stuckey, Edith B. Thompson. 1987. Spread, Impact, and Control of
Purple Loosestrife (Lythrum salicaria) in North American Wetlands. U.S. Fish
and Wildlife Service. 55 pages.
Lech, M. A. (1996). "Germination of macrophytes from a Delaware River tidal
freshwater wetland." Bulletin of the Torrey Botanical Club 123: 48-67.
Lee, H. I. and J. W. Chapman (2001). Nonindigenous species - an emerging issue for the
EPA. A landscape in transition: effects of invasive species on ecosystems, human
health, and EPA goals. U. S. E. P. Agency. Volume 2.
Lewis, W. H. and M. F. Elvin-Lewis (1977). Medical botany: plants affecting man's
health. New York, Wiley & Sons, Inc.
Lindau, C. W . and L. R. Hossner (1981). "Substrate characterization of an experimental
marsh and three natural marshes." Soil Sci. Soc. Am. J. 45(6): 1171-1176.

46

Lindig-Cisneros, R. and J. B. Zedler (2002). "Phalaris arundinacea seedling
establishment: effects of canopy complexity in fen, mesocosm, and restoration
experiments" Canadian Journal of Botany 80: 617-624.
Locandro, R. R. (1973). Reproduction Ecology of Polygonurn cuspidaturn. Department of
Botany. New Brunswick, New Jersey, USA, Rutgers University. Ph.D.
Locandro, R. R. (1978). "Weed watch: Japanese bamboo." Weeds Today 9: 21-22.
Maguire, C. E. (1985). Wetland replacement evaluation. U. S. A. C. o. Engineers.
Norfolk, Virginia.
Mal, T. K., J. Lovett-Doust, et al. (1997). "Time dependent competitive displacement of
Typha angustifolia by Lythrurn salicaria." Oikos 79: 26-33.
Mal, T. K., J. Lovett-Doust, et al. (1992). "The biology of Canadian weeds, Lythrurn
salicaria." Canadian Journal of Plant Science 72: 1305-1330.
Malakoff, D. (1998). "Restored wetlands flunk real-world test." Science 280: 371-372.
Marten, G. C. (1973). "Alkaloids in reed canarygrass." Weed Science 13: 15-31.
Maruta, E. (1976). "Seedling establishment of Polygonurn cuspidaturn on Mt. Fuji."
Japanese Journal of Ecology 26: 101-105.
Maruta, E. (1981). "Size structure in Polygonurn cuspidaturn on Mt. Fuji." Japanese
Journal of Ecology 31: 441-445.
Maurer, D. A. and J. B. Zedler (2002). "Differential invasion of a wetland grass
explained by tests of nutrients and light availability on establishment and clonal
growth." Oecologia 131: 279-288.
Merigliano, M. and P. Lesica (1998). The native status of reed canarygrass in the Inland
Northwest, USA. Abstract G 19, Final Program and Abstracts, Ecosystem
Restoration: Turning the Tide. Society for Ecological Restoration Northwest
Chapter Conference, Tacoma, Washington.
Merigliano, M. F. and P. Lesica (1998). "The native status of reed canarygrass in the
Inland Northwest, USA." Natural Areas Journal 18: 223-230.
Miller, R. C. and J. Zedler (2003). "Reponses of native and invasive wetland plants to
hydroperiod and water depth." Plant Ecology 167: 57-69.
Mitchell, D. S. and B. Gopal (1991). Invasion of tropical freshwaters by alien aquatic
plants. In P.S. Ramakrishnan , ed. Ecology of Biological Invasion in the Tropics.
New Delhi, International Scientific Publications.

47

Mitchell, E. (1926). "Germination of seeds of plants native to Dutchess County, New
York." Botany Gazette 81: 108-112.
Mitsch, W . J. and N. Flanagan (1997). Comparison of structure and function of reference
fresh-water marshes with constructed deepwater marshes at the Des Plaines River
Wetland Demonstration Project in northeastern illinois. T. O. S. U. R.
Foundation. Chicago, illinois, Wetlands Research, Inc.
Mitsch, W . J. and J. G. Gosselink (2000). Wetlands, Third Edition. New York, John
Wiley & Sons, Inc.
Morrison, S. L. and J. Molofsky (1998). "Effects of genotypes, soil moisture, and
competition on the growth of an invasive grass, Phalaris arundinacea." Canadian
Journal of Botany 76: 1939-1946.
Morrison, S. L. and J. Molofsky (1999). "Environmental and genetic effects on the early
survival and growth of the invasive grass Phalaris arundinacea." Canadian
Journal of Botany 77: 1447-1453.
Morse, R. and R. Palmer (1925). British Weeds: Their Identification and Control.
London, Ernest Benn Limited.
Muenscher, W. C. (1955). Weeds. Ithaca, New York, Cornell University Press.
Nilsson, S. G. and 1. N. Nilsson (1978). "Species richness and dispersal of vascular plants
to islands in Lake Mockeln, southern Sweden." Ecology 59: 473-480.
Niswander, S. F. and W. J. Mitsch (1995). "Functional analysis of a two-year-old created
instream wetland: Hydrology, phosphorus retention, and vegetation survival and
growth." Wetlands 15(3): 212-225.
Odland, A. (2002). "Patterns in the secondary succession of a Carex vesicaria L. wetland
following a permanent drawdown." Aquatic Botany 74: 233-244.
Ohwi, J. (1965). Flora of Japan. Washington, D.C., Smithsonian Institution.
Patterson, D. T. (1976). "The history and distribution of five exotic weeds in North
Carolina." Castanea 41: 177-180.
Pauly, W. R. (1986). "Summary of Mexican bamboo control methods (Wisconsin)."
Restoration and Management Notes 4(1): 37-38.
Pearsall, W. H. (1918). "The Aquatic and Marsh Vegetation of Esthwaite Water." Journal
of Ecology 6(1): 53-74.

48

Pojar, J. and A. MacKinnon, Eds. (1994). Plants of Costal British Columbia including
Washington, Oregon & Alaska. Vancouver, British Columbia, Lone Pine.
Potash, L. (2001). Riparian habitat restoration: Control of Japanese knotweed on the Mt.
Baker-Snoqualmie National Forest. U. S. F. S. U.S. Department of Agriculture,
Mt. Baker-Snoqualmie National Forest, Mountlake Terrace, Washington.
Pridham, A. M. S. and A. Bing (1975). "Japanese-bamboo plants." Gard 31: 56-57.
Pyek, P., B.
et a1. (2001). Persistence of stout clonal herbs as invaders in the
landscape: a field test of historical records. Plant invasions: species ecology and
ecosystem management. G. Brundu, J. Brock, 1. Camarda, L. Child and M . Wade.
Leiden, Netherlands, Backhuys: 235-244.
Pyne, S. J. (2001). Fire: A Brief History Seattle, Washington, University of Washington
Press.
Pysek, P. (1997). Clonality and plant invasions: can a trait make a
difference? The ecology and evolution of clonal plants. H. de Kroon and J. van
Groenendael. Leiden, the Netherlands, Backhuys Publishers.
Quammen, M. L. (1986). "Measuring the success of wetlands mitigation." National
Wetlands Newsletter 8: 6-8.
Race, M. S. (1986). "Wetlands restoration and mitigation policies: reply." Environmental
Management 10: 571 -572.
Rawinski, T. J. (1982). The ecology and management of purple loosestrife (Lythrum
salicaria L.) in central New York. Ithaca, New York, Cornell University. M.S.:
88.
Rawinski, T. J. and R. A. Malecki (1984). "Ecological relationships among purple
loosestrife, cattail and wildlife at the Montezuma National Wildlife Refuge." New
York Fish and Game Journal 31: 81-87.
Reimold, R. J. and S. A. Cobler (1985). Wetlands mitigation effectiveness. U. S. E . P.
Agency. Boston, Massachusetts.
Reinhartz, J. A. and E. L. Warne (1993). "Developmentof vegetation in small created
wetlands in southeastern Wisconsin." Wetlands 13: 153-164.
Ridley, H. N. (1930). The Dispersal of Plants Throughout the World. Kent, England, L.
Reeve & Company Limited.
Roberts, L. (1993). "Wetlands trading is a loser's game, say ecologists." Science 260:
1890-1892.

49


Rubio, G. and R. S. Lavado (1999). "Acquisition and allocation of resources in two
waterlogging-tolerant grasses." New Phytologist 143: 539-546.
Rudkin, W. H. (1879). "Lythrum Salicaria." Bulletin of the Torrey Botanical Club 6(54):
323.
Schnitzler, A. and S. Muller (1998). "Ecologie et biogeographie de plantes hautement
invasives en Europe: les renouees geantes du Japon (Fallopia japonica et F.
sachalinensis)." la Terre et la vie 53: 3-38.
Schoth, H. A. (1938). "Reed Canarygrass." USDA Farmer's Bulletin 1602: 11.
Seiger, L. (1995). Element stewardship abstract: Polygonum cuspidatum. Arlington,
Virginia, Nature Conservancy.
Shamsi, S. R. A. and F. H. Whitehead (1974). "Comparative Eco-Physiology of
Epilobium hirsutum L. and Lythrum salicaria L. :1. General Biology, Distribution
and Germination." Journal of Ecology 62(1): 279-290.
Shamsi, S. R. A. and F. H. Whitehead (1977). "Comparative eco-physiology of
Epilobium hirsutum L. and Lythrum salicaria L. III. mineral nutrition." Journal of
Ecology 65: 55-70.
Sheaffer, C. C., P. R. Peterson, et al. (1992). "Drought effects on yield and quality of
perennial grasses in the north central United States." J. Prod. Agric. 5: 556-561.
Sifneos, J. c., J. E.W. Cake, et al. (1992). "Effects of section 404 permitting on
freshwater wetlands in Louisiana, Alabama, and Mississippi." Wetlands 12: 28­
36.
Skinner, L., W. J. Rendall, et al. (1994). Minnesota's Purple Loosestrife Program:
History, Findings, and Management Recommendations., Minnesota Department
of Natural Resources; Special Publication: 1-25.
SolI, J. and J. Morgan (Undated). Japanese Knotweed, The Nature Conservancy of
Oregon.
Spuhler, D. R. (1994) . Low plant diversity found in communities dominated by Reed
Canary Grass (Phalaris arundinacea). Fourteenth Annual North American Prairie
Conference.
Stannard, M. and W. Crowder (2001). Biology, History and Suppression of Reed
Canarygrass (Phalaris arundinacea L.). Boise, Idaho.

50


States, S. C. o. t. U. (2002). Borden Ranch Partnership and Angelo K. Tsakopoulos v.
United States Army Corps of Engineers and Environmental Protection Agency.
Washington, D.C., Department of Justice.
Stout, A. B. (1925). "Studies of Lythrum Salicaria - II* A new form of flower in the
species." Bulletin of the Torrey Botanical Club 52(3): 81-85.
Streever, B. (1999). "Examples of performance standards for wetland creation and
restoration in Section 404 permits and an approach to developing performance
standards." WRP Technical Notes Collection TN WRP WG-RS-3.3: U.S. Army
Engineer Research and Development Center, Vicksburg, MS.
Stucky, R. L. (1980). "Distributional History of Lythrum salicaria (purple loosestrife) in
North America." Bartonia 47: 3-20.
Svengsouk, L. and W. J. Mitsch (1997). Five-year monitoring of a mitigation wetland in
Central Ohio. The Olentangy River Wetland Research Park at The Ohio State
University, 1996 Annual Report. W. J. Mitsch. Columbus, Ohio, School of
Natural Resources, The Ohio State University.
Swink, F. A. and G. S. Wilhelm (1994). Plants of the Chicago Region, Fourth Edition.
Indianapolis, Indiana, Indiana Academy of Science.
Terzi, G. (2006). Invasive Species - An Evolving Policy, U.S. Army Corps of Engineers,
Seattle District, Regulatory Branch.
Thompson, D. Q., R. L. Stuckey, et al. (1987). Spread, Impact, and Control of Purple
Loosestrife (Lythrum salicaria) in North American Wetlands. U. S. F. a. W.
Service. Jamestown, ND, Northern Prairie Wildlife Research Center: 55.
Tong, T. (2006). Email. C. Ehorn. Seattle, Washington.
Torrey, J. (1877). "Lythrum Salicaria." Bulletin of the Torrey Botanical Club 6(32): 171.
Torrey, R. H. (1931). "Field Trips of the Club." Torreya, The Torrey Botanical Club 31:
16-18.
Townsend, A. (1997). "Japanese knotweed: a reputation lost." Arnoldia 57: 13-19.
Toxicology, B. o. E. S. a. (2001). Compensating for Wetland Losses Under the Clean
Water Act Washington, D.C., National Academies Press.
Tsvelev, N. N. (1983). Grasses of the Soviet Union, Part 1. New Delhi, Oxonian Press
Private Limited.

51


Turner, N. J. (1992). "'Just when the wild roses bloom' : The legacy of a Lillooet basket
weaver." TEK TALK: A Newsletter of Traditional Ecological Knowledge 1: 2-5.
Twolan-Strutt, L. and A Keddy (1996). "Above- and belowground competition intensity
in two contrasting wetland plant communities." Ecology 77: 259-270.
U.S. Department of Agriculture, N. R. C. S. (1996). Draft reed canarygrass control
project plan. P. M. Centers, Corvallis, Oregon and Pullman, Washington.
Voegtlin, D. (1998). Biological Control of Purple Loosestrife Program. University of
Illinois, Urbana-Champagne, Center for Ecological Entomology.
Wang, S., Z. Zheng, et al. (2004). "Angiogenesis and anti-angiogenesis activity of
Chinese medicinal herbal extracts." Life Science 74(20): 2467-2478.
Wasser, C. H. (1982). Ecology and culture of selected species useful in revegetating
disturbed lands in the West. U. S. F. a. W. Service.
Weiher, E., 1. C. Wisheu, et al. (1996). "Establishment, persistence, and management
implications of experimental wetland plant communities." Wetlands 16: 208-218.
Weiher, P. E. and R. K. Neely (1997). "The effects of shading on competition between
purple loosestrife and broad-leaved cattail." Aquatic Botany 59: 127-138.
Wetzel, P. R. and A G. van der Valk (1998). "Effects of nutrient and soil moisture on
competition between Carex stricta, Phalaris arundinacea, and Typha latifolia."
Plant Ecology 138: 179-190.
Wheeler, W. A (1950). Forage and Pasture Crops: a handbook of information about the
grasses and legumes grown for forage in the United States. New York, New York,
Van Nostrand Company.
Wilcove, D. S., D. Rothstein, et al. (1998). "Quantifying threats to imperiled species in
the United States." Bioscience 48: 607-615.
Wilcox, D., A (1989). "Migration and Control of Purple Loosestrife (Lythrum salicaria
L.) along Highway Corridors." Environmental Management 13(3): 365-370.
Wilcox, D. A, M. K. Seeling, et al. (1988). Ecology and management of potential for
purple loosestrife (Lythrum salicaria). Management of exotic species in natural
communities. L. K. Thomas. Fort Collins, Colorado, Colorado State University.
Volume 5.
Wilkins, F. S. and H. D. Hughes (1932). "Agronomic trials with reed canary grass."
Journal of the American Society of Agronomy 24: 18-28.

52

Wilson, R. F. and W. J. Mitsch (1996). "Functional Assessment of five wetlands
constructed to mitigate wetland loss in Ohio, USA." Wetlands 16: 436-451.
Yoshioka, K., Ed. (1974). Volcanic vegetation. Flora and Vegetation of Japan.
Amsterdam, Elsevier.
Zedler, J. B. (1996). "Costal mitigation in southern California: The need for a regional
restoration strategy." Ecological Applications 6: 84-93.
Zeiders, K. E. (1976). "A new disease ofreed canarygrass caused by Helminasporium
catenarium." Plant Disease Reporter 60: 556-568.

53

Media

Ehorn_CHMESThesis2006.pdf