Fire Effects Information System (FEIS)
FEIS Home Page

Phragmites australis


Photo ©2006 Susan Vincent

Gucker, Corey L. 2008. Phragmites australis. In: Fire Effects Information System, [Online]. U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station, Fire Sciences Laboratory (Producer). Available: /database/feis/plants/graminoid/phraus/all.html [].



common reed
Danube grass
Roseau cane

The scientific name of common reed is Phragmites australis (Cav.) Trin. ex Steud. (Poaceae) [14,58,72,111,126]. Common reed belongs to the Panicoideae subfamily and the Arundineae tribe [58].

Currently a single subspecies and variety are recognized:

Phragmites australis subsp. americanus Saltonstall, PM Peterson & Soreng [197], native lineage
Phragmites australis var. berlandieri (E Fourn.) CF Reed [197], Gulf Coast lineage or haplotype I

Recent and previously uncharacteristic increases in common reed abundance led to the study of its genetics. Saltonstall [196] determined that 11 native haplotypes and 1 introduced haplotype occur throughout North America. The introduced haplotype (M) is of European origin and is referred to as the "nonnative haplotype" throughout this review.

Phragmites communis Trin. [112,158,190,215,249]
Phragmites communis var. berlandieri (Fourn.) Fern [190]
Phragmites phragmites ( L.) Karst. [194,246]


No special status

Information on state-level noxious weed status of plants in the United States is available through Plants Database.


SPECIES: Phragmites australis
Common reed is one of the most widely distributed flowering plants [15,114]. It occurs on every continent except Antarctica [190] and is cosmopolitan in temperate zones [136]. Common reed is widely distributed in North America, occurs in all US states except Alaska, and in all Canadian provinces and territories except Nunavut and Yukon [112]. Common reed is native to Puerto Rico and occurs as a nonnative in Hawaii [73,231]. Grass Manual on the Web provides a map of common reed's North American distribution.

Subspecies, variety, and haplotype distributions: Extensive genetics studies on common reed plant material from modern and herbarium samples (dated to the 1850s) collected throughout North America revealed there are 11 native haplotypes and 1 nonnative haplotype [196]. There were significant changes in common reed haplotype frequencies between historic (herbarium samples collected pre-1910) and modern samples (P<0.001). Introduction of the nonnative haplotype probably occurred at 1 or more Atlantic Coast ports early in the 19th century, and because morphological differences between the haplotypes are subtle, the introduction(s) went unnoticed. Range expansion of the nonnative haplotype was likely facilitated by travel way construction during this time period [195]. The nonnative haplotype is dominant along the Atlantic Coast and in the Great Lakes area. In western North America, the nonnative haplotype is becoming common along roadsides and waterways in urban areas, but native types are still common in the Southwest and Pacific Northwest [196].

P. australis subsp. americanus is native to the United States. Its current range extends from the southwestern Northwest Territories south to southern California, east to northern Texas, northern Arkansas, North Carolina, West Virginia, and north to Newfoundland and Quebec [197].

P. australis var. berlandieri may or may not be native to North America, but if introduced was a much earlier introduction than the nonnative haplotype. The current distribution of P. australis var. berlandieri is not different from historic distributions [196]. Phragmites australis var. berlandieri, also known as the Gulf Coast lineage, occurs along the Gulf Coast of Mexico, in South America, and on the Southern Pacific Islands [195]. In the United States, P. australis var. berlandieri occupies southern habitats from California east to Florida [14,197].

The nonnative common reed haplotype is widely distributed in North America. It occurs from British Columbia east to Quebec and south throughout the contiguous United States [14,197].

Since its introduction, the nonnative haplotype has expanded its range throughout North America and most dramatically along the Atlantic Coast and in the Great Lakes area. The nonnative type replaced native types in New England and established in the southeastern United States, where native common reeds did not occur historically. In Connecticut and Massachusetts, 19th century common reed samples were primarily native haplotypes, but by 1940, all samples were nonnative. Local extinctions of native haplotypes are not uncommon [195]. In Falmouth, Massachusetts, researchers located 268 common reed populations; 4 were native [175]. Native and nonnative common reed populations were mapped for all of Rhode Island; native populations were restricted to the eastern side of Block Island, and the largest stand was about 2 acres (1 ha) [137,139]. On Delmarva Peninsula, Maryland, nonnative common reed is most common, but the average size of nonnative populations is often much smaller than that of native populations [161].

In Quebec, the nonnative haplotype was present as early as 1916 but was rare before the 1970s and restricted to shores of the St Lawrence River. In less than 20 years, the nonnative haplotype became dominant; over 95% of colonies sampled were nonnative [146]. In semiurban landscapes of southern Quebec, the nonnative common reed haplotype was most common in linear wetlands, industrial areas, and rights of way. Intrinsic rates of increase (r) in these areas were determined using a nonlinear growth model that compared clone size at time zero to the clone size years after the initial observation. In St-Bruno-de-Montarville, the intrinsic rate of increase ranged from 0.19 to 0.34/year. On the east tip of Laval Island, the intrinsic rate of increase ranged from 0.19 to 0.54/year. Riparian habitats had less common reed than anthropogenic wetlands. The number of colonization events at rights of way was high. For a discussion on the possible role of colonization by seed, see Seed production [154].

Changes in local distributions: General increases in the area occupied by common reed have been reported in many places; however in some cases, nativity of the population is not identified. Establishment and spread patterns may vary with degree of anthropogenic disturbance, haplotype, salinity levels, and stand age. Additional information is available in the sections on Regeneration Processes and Successional Status.

In a review, Chambers and others [43] found that early reports of common reed abundance described it as "occasional," "not common," or "rare". By the late 1990s, common reed was described as a "widespread" "nuisance species". Increases in common reed abundance in these areas generally coincided with increased human manipulation of coastal areas and wetlands [43]. Aerial photos taken from 1955 to 2000 showed that the area dominated by common reed between 1995 and 1999 increased exponentially on Long Point, southwestern Ontario. Of the 31 common reed stands that were sampled in or after 2000, 90% were nonnative. Researchers suggested that establishment and spread of the nonnative type was the primary reason for increased dominance, and suggested that increased temperatures and decreased water levels in the mid- to late 1990s may have favored increased spread [252].

Local increases in common reed are reported from several areas, although nativity of the populations is unknown. On the Tailhandier Flats on Quebec's St Lawrence River, common reed increased the surface area occupied by 18% from 1980 to 2002 based on aerial photos and remote sensing data [117]. In central Washington, aerial photos of the Winchester Wasteway showed that the area occupied by common reed increased 39 acres (15.8 ha) in 3 years [115]. Researchers compared time series maps to track the establishment and spread of common reed populations in mid-Atlantic coastal areas. Spread rate averaged 10 acres (5 ha)/year. Area occupied by common reed increased rapidly up to 20% per year until stands covered 50% to 80% of a given marsh. Patchiness was common soon after establishment but decreased over time. Common reed abundance decreased at only one site, Lang Tract, Delaware, and decreases were temporary. In southwestern Louisiana's Rockefeller Wildlife Refuge, the size and number of common reed clones increased over time after its introduction in 1968. Estimated intrinsic rates of increase of 21 common reed clones ranged from 0.0767 to 0.2312/year. Lag time between establishment and rapid expansion was 10 to 15 years [212].

Common reed is widespread in both estuarine intertidal and palustrine persistent emergent wetlands [49]. It often forms monotypic stands [10,94], as other species are excluded by persistent shading and extensive utilization of space by common reed [100].

Although common reed stands are often monotypic, adjacent wetter and drier sites may be occupied by more flood-tolerant and less flood-tolerant species, respectively [94]. Dominant vegetation within a wetland or riparian site is often determined by water levels and flood tolerances, and so it often fluctuates with water table changes [225]. These zones of vegetation are "extensive and dramatic" in Big Creek Fen of Cherry County, Nebraska [32], and well-defined in swamps of northwestern Minnesota [66]. In the Delta Marshes of southern Manitoba, it appears that common reed is the only species for acres, but a closer look reveals patches of common river grass (Scolochloa festucacea) within the stands [150]. Disturbances can also affect community composition. In southern and eastern Idaho and eastern Montana, nonnative Canada thistle (Cirsium arvense) may establish in highly disturbed common reed stands [92,95].

Common reed is a dominant species in the following vegetation types and classifications recognized in the United States and Canada. Broad classifications are presented before state-specific classifications.

Throughout the United States:

Rocky Mountains:

Great Plains:

Canadian Prairie Provinces:

Southern United States:











New York:





SPECIES: Phragmites australis
This description provides characteristics that may be relevant to fire ecology, and is not meant for identification. Keys for identification are available (e.g., [58,82,87,111,158,190,215]).

Phragmites australis subsp. americanus, P. a. var. berlandieri, and the nonnative common reed haplotype are distinguished morphologically by the Flora of North America [14] and Blossey [26]. As new information is available, discriminating morphological characteristics are updated at [26].

Aboveground description: Common reed is a robust perennial grass that may reach 20 feet (6 m) tall [84,127,215]. It is the tallest native grass in Nova Scotia [190], Montana [136], and possibly other states or provinces. Maximum height is not typically reached until plants are 5 to 8 years old [52]. Common reed spreads by clonal growth via stolons and rhizomes, and produces dense stands [51,85,111,127]. Clones are long-lived; some report clones may persist for over 1,000 years (Rudescu and others 1965, cited in [100]), but no portion of the clone lives more than 8 years. Rhizomes typically outlive aboveground shoots [102]. Stolons are most typical during times of low water and reach lengths of up to 43 feet (13 m) [142,235].

Common reed produces stout, erect, hollow aerial stems [169,181]. Stems are usually leafy, persistent, and without branches [15,247]. At the base, stem thickness measures 5 to 15 mm [15,142]. Leaves are aligned on one side of the stem, flat at maturity, and measure 4 to 20 inches (10-60 cm) long and 0.4 to 2 inches (1-6 cm) wide [58,87,112,159]. Leaf margins are somewhat rough [85], and leaves are generally deciduous [111]. Common reed flowers occur in a large, feathery, 6- to 20-inch (15-50 cm) long panicle [63,181]. The panicle has many branches and is densely flowered [159]. Panicles are up to 8 inches (20 cm) wide after anthesis [82]. Spikelets contain 1 to 10 florets. Floret size decreases from the base of the panicle upward. Lower florets are staminate or sterile and without awns. Upper florets are pistillate or perfect with awns. Occasionally all spikelets are abortive [46,87,111,142,247]. Sometimes spikelets are reduced to a single glume and floret, causing panicles to lose their feathery appearance [235]. Seeds are small, measuring up to 1.5 mm long [142]. Common reed seeds collected from a salt marsh near the mouth of Delaware Bay had an average air-dry mass of 125.2 µg [251].



Photos ©Gary Fewless
Cofrin Center for Biodiversity
University of Wisconsin-Green Bay

Belowground description: Extensive rhizome and stolon growth produces dense common reed stands [51,85,111,127]. First-year common reed rhizomes observed in Britain typically produced only 1 aboveground stem. In the 2nd year, rhizomes produced up to 4 aboveground stems, and in the 3rd year rhizomes produced up to 6 aboveground stems. Stem production usually decreased after rhizomes reached 6 years old [99].

Rhizomes are thick, "deep seated", and scaly [142,159] and can grow to 70 feet (20 m) long [114]. Rhizomes may grow 16 inches (40 cm)/year [54] and live 2 to 3 years [114]. Rhizomes in soil are commonly long, thick, and unbranched. In water, rhizomes are more slender, produce multiple branches, and are often shorter [114]. In the Prairie Provinces, common reed plants growing in wet soil at the water's edge produced thick, soft, spongy rhizomes that branched in several directions and at several levels. There were clusters of roots bearing other hair-like roots at the nodes [107].

Common reed rhizomes can penetrate deeply, but rhizome depth varies with site conditions. On the Atlantic coast of Delaware, researchers described common reed's belowground growth as a thick rhizome mat 4 to 8 inches (10-20 cm) below the surface [80]. In swamps of Cherry County, Nebraska, common reed rhizomes were 30 feet (9 m) deep [225]. An "extraordinarily large number" of rhizomes and roots formed a dense mat from the soil surface to about 8.2 feet (2.5 m) deep in the Skokie Marsh of Illinois [203]. In the Riverbend Marsh area of New Jersey's Hackensack Meadowlands, common reed roots and rhizomes in interior high marshes reached 24 inches (60 cm) deep and in mosquito ditches reached 22 inches (55 cm) deep [19]. Depth of belowground structures averaged 9.8 inches (25 cm) in clay soils and averaged 16 inches (40 cm) in moister soils with lower clay content on the southern coast of New Hampshire [36]. Additional information on rhizome, stolon, and clonal growth is available in Vegetative regeneration.


Common reed reproduces sexually from seed and vegetatively from stolons and rhizomes.

Local spread of common reed is predominantly through vegetative growth and regeneration, while establishment of new populations occurs through dispersal of seeds, rhizomes, and sod fragments. For example, on the Tailhandier Flats on Quebec's St Lawrence River, common reed increased its surface area occupied by 18% from 1980 to 2002. Researchers attributed an average of 88% of the spread to vegetative growth but suggested that new colonies were the result of seedling establishment [117]. Near the mouth of Delaware Bay, common reed moved into salt marshes through rhizome and stolon growth from more upland sites. Establishment from seed occurred in sparsely populated or bare patches within the marsh. Some bare site colonization may occur through vegetative growth, but vegetative colonization likely decreases as distance from an established population increases [251].

Reproductive mode affects the genetic makeup of common reed populations. In the Charles River Watershed of Massachusetts, the genetic makeup of clones that made up stands and stands that made up populations were evaluated. Stands were mosaics of different clones. Populations were closely related, but plants within populations were more closely related than plants from different populations. The researcher concluded that colonization was likely vegetative, and populations increased over a short time period [129].

Breeding system: Common reed produces male, female, and perfect flowers. Lower florets are staminate or sterile, and upper florets are pistillate or perfect [87,247].

Pollination: Cross pollination of common reed flowers is probably most common, but self pollination or agamospermy (seed production without fertilization) are also possible. In the laboratory, 5 of 16 native inflorescences and 2 of 4 nonnative inflorescences from populations in Rhode Island produced viable seed through either self pollination or agamospermy [138]. Some self pollination also occurred in common reed populations in Japan, although seed set was much lower for self-pollinated than cross-pollinated flowers [119].

Seed production: Many researchers indicate that common reed rarely produces viable seed [82,97,235], while others indicate that viable seed is produced at least sometimes in some locations. Voss [235] reported that "fertile seed is often not developed, (and common reed) customarily reproduces vegetatively". In Colorado, some common reed populations produced empty spikelets and were likely limited to vegetative regeneration [242]. Some researchers indicated that early frosts in the Delta Marsh of south-central Manitoba prevent successful seed production [150]; however, Shay and Shay [202] reported viable seed production in the Delta Marsh and observed seedlings on drying shorelines in the area. Ailstock (unpublished data, cited in [3]) reported that overwintering common reed inflorescences produced abundant viable seed. Common reed plants growing near the mouth of Delaware Bay produced 500 to 2,000 seeds/shoot [251]. Seed set averaged 9.7% and ranged from 0.1% to 59.6% for 12 common reed populations in southwestern Japan. Flowers from 2 cross-pollinated populations set 52.4% and 64.4% of seed. Self-pollinated flowers produced 2.8% and 8.9% of seed [119]. From common reed populations in St-Bruno-de-Montarville, Quebec, an average of 6.6% and a maximum of 27.1% of seeds were viable. From populations on the east tip of Laval Island, Quebec, an average of 2.7% and a maximum of 11.3% of seeds were viable. Based on the abundance of flowers produced/inflorescence, researchers estimated 350 to 800 viable seeds could be produced/inflorescence [154].

Viable seed production may be affected by site factors, but there is little information on the conditions necessary for successful common reed seed development. According to Cross and Fleming [52], common reed may need to reach 3 or 4 years old before producing viable seed. In Utah's Fish Spring National Wildlife Refuge, there are 2 distinct common reed communities. A dwarf community with limited rhizome growth occurring between greasewood (Sarcobatus vermiculatus) and saltgrass vegetation may have established from seed. Within the marsh, common reed has substantial vegetative growth [30].

Seed dispersal: Common reed seeds are dispersed by wind [251] and water. Buoyancy of seeds from Germany and the Netherlands may be slightly less in stagnant than moving water. Ninety percent of seeds were still floating after 10 days in stagnant water and after 23 days in moving water. Half of seeds were floating after 32 days in stagnant and after 69 days in moving water, and 10% of seeds were still floating after 121 days in stagnant water and 124 days in moving water [232].

On salt hay (saltmeadow cordgrass (Spartina patens), saltgrass (Distichlis spicata), and/or saltmeadow rush (Juncus gerardii)) farms in Commercial Township, New Jersey, common reed established only after Hurricane Hazel in 1954. It is likely that establishment occurred by seed brought by storm tides from Delaware. However, vegetative propagules may have also been carried in the storm [18]. Dispersal of vegetative propagules is common in some situations. For more information, see Vegetative dispersal.

Seed banking: Information on common reed seed banking is sparse; however, several studies report some common reed seedling emergence from soil seed banks. Although submersion often reduces emergence, it does not necessarily cause an immediate loss of viability [47,206].

Some studies and researchers indicate that common reed seed banks are small and/or short-lived [59,110]. A review by DiTomaso and others [59] reports that common reed seeds are short-lived under field conditions and that persistent seed banks are not produced. In wetlands of the Great Lakes area, common reed was present in the aboveground vegetation of all sites sampled, but no seedlings emerged from collected soils [110]. No common reed seedlings emerged from soil samples collected from back dunes of the Cape Cod National Seashore in Massachusetts, but common reed was rare in the study area [13]. Common reed did not emerge from soils collected in July from marshes on Wisconsin's Green Bay where its relative abundance was up to 4.1%. Soil samples were collected before seed set in the current year in order to characterize the persistent seed bank [74].

Several studies report common reed emergence from soils collected in various communities, and emergence was usually greatest from unflooded soils collected in common reed vegetation. Twenty-five soil samples were collected in early April from 6 vegetation types in Utah's Ogden Bay Waterfowl Management Area. Sixty-four common reed seedlings/m² emerged from soil collected in common reed stands. From soil collected in hardstem bulrush (Scirpus acutus) and cattail (Typha spp.) stands, 2 and 4 common reed seedlings/m² emerged, respectively. There was no common reed emergence from soil collected in the other 3 vegetation types. When soil samples were submerged, no common reed seedlings emerged. Researchers noted that common reed emergence was low compared to other emergent vegetation, and there were no common reed seedlings on an unvegetated, recently drained mudflat in the study area [206]. Common reed seedlings emerged from soils collected in June from 8 cover types in the Delta Marsh. Seedling density was lowest (5 seedlings/m²) in soils collected from large bays and greatest (90 seedlings/m²) in soils from common reed-dominated vegetation. Large bays often had less than 3 feet (1 m) of standing water. Submergence of soils in the greenhouse also affected emergence; 398 common reed seedlings emerged in drawdown and 4 emerged in submerged (0.8-1.2 inches (2-3 cm)) conditions [177].

Common reed seeds can survive submergence. Emergence generally decreases in flooded conditions, but a short period of submersion may increase germination success. Of common reed seeds submerged in 12 inches (30 cm) of water in a canal in Prosser, Washington, 16%, 51%, and 54% germinated after 3, 6, and 9 months of submergence, respectively. Germination decreased to 5% and 1% germinated after 36 and 60 months of submergence, respectively. Mature seeds were collected in the field and stored for 1 year at room temperature before submergence treatments. After 60 months of dry storage and no submergence, 16% of common reed seed germinated [47]. Common reed seedlings established on a mudflat on northwestern Minnesota's Mud Lake National Wildlife Refuge, but seeds collected nearby did not germinate after wet, outdoor storage treatments. After 7 months of dry storage at room temperature, about half of the common reed seeds germinated. Germination decreased to about 30% after 8 months of dry, room temperature storage [98].

Germination: Warm temperatures, high light conditions, and low to moderate salinity levels on moist but not flooded sites are most conducive to successful common reed seed germination.

Stratification for 6 months at 39 °F (4 °C) was required for germination of common reed seed collected from the Delta Marsh. Under full light, all seeds germinated at alternating temperatures of 68 °F (20 °C) and 86 °F (30 °C) and 97% germinated at alternating temperatures of 59 °F (15 °C) and 77 °F (25 °C) [79].

Germination of common reed seed collected from the Delta Marsh was best on the soil surface in full light. The maximum germination rate was 52% in dark conditions. In full light, germination rates were 70% on the soil surface, 30% at 0.4 inch (1 cm) deep, and 12% at 1.6 inches (4 cm) deep. No common reed seedlings emerged when seeds were buried 2 inches (5 cm) deep. Optimal germination temperatures were maintained during burial experiments [79].

Common reed germination may be decreased at salinity levels greater than 5,000 ppm [36]. Seed from Meadow Pond and Little River Salt Marsh on the southern New Hampshire coast germinated at 35.5% in fresh water, 36.5% at 5,000 ppm salinity, and 11% at 20,000 ppm salinity. A single seed germinated at 30,000 ppm salinity, and no seed germinated at 35,000 ppm salinity [36]. Another study showed similar results, with 4% of seeds germinating in a salt-free environment, 36% at 2,000 ppm salinity, and 32% at 5,000 ppm salinity [79].

Oxygen is required for common reed germination; however, exposure to anoxic and high-salt conditions may increase germination once seeds are returned to salt-free environments and atmospheric oxygen levels. Without oxygen, common reed seeds collected near the mouth of Delaware Bay in November did not germinate in any salinity level from 0 to 40,000 ppm. At atmospheric oxygen levels, germination of common reed was reduced and inhibited at 25,000 and 40,000 ppm salinity, respectively. Germination increased with salinity levels of 5,000 and 10,000 ppm when oxygen levels were reduced to 5% and 10%. Seeds treated to high salinity levels and anoxic conditions had 60% germination (maximum for the study) when returned to atmospheric oxygen levels and fresh water [250,251].

Seedling establishment/growth: Common reed establishment from seed occurs on some sites [98,245], but mortality rates are high when seedlings are exposed to flooding, drought, salt, and freezing [52,102], (Hurlimann 1951, cited in [101]). Seedling sensitivities may limit establishment from seed to ideal site and weather conditions.

Common reed seedling survival is often low. In Stemmers Run Wildlife Management Area in Maryland, common reed established from seed on bare high marsh soils, but after 12 weeks survival was just 0.7%. Survival of seeds collected and grown in a greenhouse was 27% [3]. Research by Haslam [104] indicates that winter mortality is high for young common reed plants with only 1 to 3 shoots and no rhizome development and is low for plants with 10 to 12 shoots. Common reed seedlings growing for 2 to 4 seasons can have just 3 shoots and no horizontal rhizome growth or may have over 200 shoots, be up to 4.3 feet (1.3 m) tall, and occupy an area over 22 ft² (2 m²) [104].

Common reed seedling establishment is typically restricted to muddy sites with "just enough water". High water levels can drown or wash away seedlings, and too little moisture leads to desiccation. Once seedlings reach 5 to 6 inches (13-15 cm) tall they can typically survive flooding depths of 3 to 4 inches (8-10 cm) [210].

Warm temperatures, high light levels, and high phosphate levels can provide for "good" seedling growth. Based on research conducted in England, Haslam [104] reports that seedlings grow faster at 77 °F (25 °C) than at 59 °F (15 °C). In low light, seedlings appear small and weak. When phosphate levels are low, seedling growth is stunted [104].

Field sites suitable for seedling emergence are typically unflooded and unshaded. Common reed seedlings emerged in the Delta Marsh after flooded sites were drawn down to 12 inches (30 cm) below normal. Seedling recruitment in the field was compared to emergence from soil samples. The maximum seedling density was 25 seedlings/m² from soil samples and 20 seedlings/m² in the field. In the field, the largest number of common reed seedlings occurred in the draw down area with low common reed density. Seedling recruitment was lacking in dense common reed stands, and recruitment was low on sites with 500% to 600% moisture [245]. On salt hay farms in New Jersey, common reed established after Hurricane Hazel on bare areas that were inundated the longest and on newly constructed dikes and berms. However, whether or not this was establishment from seed or vegetative propagules is unknown [18].

Native and nonnative seedling growth: Common reed seedling establishment, growth, and mortality can vary among haplotypes. Native common reed seedlings suffered higher mortality, produced less below- and aboveground biomass, and were shorter than nonnative seedlings under low- and high-nutrient treatments. Researchers compared native and nonnative growth in an outdoor experiment with seeds collected from Ontario, Rhode Island, Maryland, and Delaware. After 1 month, 23% of native and 15% of nonnative seedlings died. All native seedlings from seed collected in Rhode Island died within 2 weeks. At the end of the experiment (about 9 months), 38% of native and 23% of nonnative seedlings were dead. In the high-nutrient treatment, nonnative seedlings produced significantly more rhizome biomass (x=113.8 g) than native seedlings (x=44.3 g) (P<0.0001). Native seedling stems were clustered around where the seed was planted, but nonnative stems were spread throughout. No native seedlings flowered, but 3 nonnative seedlings did. Above- and belowground biomass and number of shoots produced by nonnative seedlings were 2 to 4 times those of native seedlings in low- and high-nutrient treatments [198].

Average above- and belowground biomass, number of shoots, and shoot height of native and nonnative common reed seedlings [198]
  Low nutrients High nutrients
Native Nonnative Native Nonnative
Aboveground biomass (g) 14.3 51.7 84.3 183.5
Belowground biomass (g) 13.3 54.9 79 155.8
Number of shoots 12.2 27 44.3 86.5
Shoot height (cm) 67.1 99.9 99.2 115.4

Vegetative regeneration: Once established, common reed regeneration and spread are primarily through rhizome and sometimes stolon growth. A substantial amount of common reed establishment also occurs vegetatively through colony breakage and dispersal of rhizome fragments [3,210]. Vegetative growth allows common reed to spread into sites unsuitable for establishment from seeds. Common reed rhizome production and vegetative spread can be extensive. For additional information on the morphology, spatial distribution, and structure of common reed rhizomes and stolons, see General Botanical Characteristics.

Vegetative dispersal: Vegetative rhizome and stolon growth is the predominant method of common reed spread following establishment [235,242], but rhizome and sod fragments also provide for successful establishment. Rhizome fragments often establish and survive better than seeds [3], and young plants produced from rhizomes are generally less sensitive than seedlings. For more on the establishment of common reed from seeds, see Seedling establishment/growth.

Common reed is often dispersed through the transport of rhizome fragments and the movement of sod. Mechanical equipment operating in a common reed-dominated community had 69 rhizome buds in its tracks (Ailstock, personal observation, cited in [3]). Rhizome fragments with 2 to 3 nodes are often viable [52]. Small portions of common reed stands can be torn from river banks, float downstream, and reestablish. In Leech Lake, northern Minnesota, an entire common reed stand was dislodged by a storm. The stand moved and reestablished about 1 mile (1.6 km) from its original location [210].

Field and greenhouse studies suggest that the survival of common reed shoots produced by rhizome fragments is better than that from seeds. Buried rhizomes survived better than seeds when both were collected from Stemmers Run Wildlife Management Area, Maryland, and grown in the greenhouse and the field. Seedling survival in the greenhouse was 27% on bare soil. Rhizomes left on the soil surface did not establish in the field. All buried rhizomes survived in vegetated, burned, and bare soils in the greenhouse. In the field, buried rhizome survival was 10%, 30%, and 20% in burned, vegetated, and bare high-marsh soils, respectively. In the field, establishment from seed occurred in areas of exposed mineral soil in the high marsh dominated by switchgrass (Panicum virgatum) and common rosemallow (Hibiscus moscheutos subsp. moscheutos). Seedling survival was low; after 12 weeks, just 0.7% of seedlings were alive [3].

Vegetative establishment: High levels of salinity (18,000 ppm), anoxic conditions, exposure, and small rhizome size can reduce the chances of successful establishment from common reed rhizome fragments [20]. Portions of common reed stands or sod may survive drought and saline conditions better than rhizome fragments [212]. Young plants established through vegetative means can be much hardier than seedlings [210].

Emergence from unburied, flooded rhizomes failed when salinity levels were high (18,000 ppm) in greenhouse and field studies in Riverbend Marsh, New Jersey. Rhizomes on the surface desiccated or washed away. No emergence occurred in poorly drained conditions, although mature common reed occupied poorly drained sites in the field [20].

Survival and shoot height were greater in fresh than saline water, and exposure to fresh water before saline water increased shoot survival and height in a greenhouse study using rhizomes and water from Riverbend Marsh. Larger rhizomes (2-node section weighing 4 g) established in saline water (9,000-21,000 ppm), but small rhizomes (2-node section < 2 g) did not. No shoots emerged from rhizomes in poorly drained or flooded treatments. A freshwater treatment before exposure to the salty Riverbend Marsh water increased shoot survival and height over all saltwater treatments. Field observations indicated that common reed established near mosquito ditches, creek banks, landfills, and railroad beds and spread vegetatively into the interior high marsh [21].

Average survival and shoot height from common reed rhizome fragments [21]
Treatment Proportion surviving to end of growing season Shoot height (cm)
Freshwater 0.80a 63.58a
Freshwater for 2 weeks, then Riverbend water (salinity 9,000-21,000 ppm) 0.67ab 45.35ab
Riverbend water 0.49b 36.39b
Values within a column followed by different letters are significantly different (P<0.05).

In areas of southwestern Louisiana's Rockefeller Wildlife Refuge, where vegetation was clipped or killed by herbicide, common reed sod established but seedlings did not. Common reed cover was greatest when sod was planted in clipped sites. Five of 10 sods survived in clipped areas, 4 of 10 survived in undisturbed areas dominated by saltgrass and saltmeadow cordgrass, and 1 of 10 survived in herbicide-killed areas. After the 2-year study period, 30% of sod pieces survived, although water tables were low and pore water salinity was 20,000 to 38,000 ppm. During the study, there were record-setting growing-season droughts. Fertilization did not affect common reed cover [212].

Nonnative common reed sprouts survived and grew better in fresh and saline environments than did native sprouts. Rhizomes were collected in Delaware and Maryland from nonnative and native haplotypes. Nonnative rhizomes produced numerous shoots. Shoots produced by native rhizomes were fewer but thicker and taller than nonnative shoots. Shoot differences persisted and were used to distinguish haplotypes 1 year later. Nonnative shoot survival was higher than native shoot survival over the salinity range tested (0-23,400 ppm, P=0.02). Native haplotypes did not grow over 2 inches (5 cm) at salinity levels above 12,870 ppm. The nonnative haplotype grew to 8 inches (20 cm) and was producing new shoots at 23,400 ppm salinity. No new native shoots were produced at salinity levels above 12,870 ppm. In freshwater, nonnative common reed produced 1.63 shoots/g dry mass of rhizome tissue, and native haplotypes produced 0.52 to 0.92 shoots/g dry rhizome tissue [233].

Survival of nonnative and native common reed haplotypes with increasing salinity [233]
Salinity (ppm) Survival (%)
Nonnative haplotype (M) Native haplotype (F) Native haplotype (AC)
1,176 80 80 80
7,605 100 20 40
12,870 100 0 20
18,720 100 0 40
23,400 100 0 0

Vegetative spread: Rhizomatous growth allows common reed to spread quickly [15] and to occupy sites unsuitable for establishment by seed or rhizome fragments [5,19]. In England's Breckland fens, common reed rhizomes grew 20 to 80 inches (50-200 cm) annually, while stolons grew to over 40 inches (100 cm) long [101]. After studying Wisconsin wetland vegetation, Curtis [54] reported that common reed rhizomes can grow 16 inches (40 cm)/year. At Horn Point marsh in the upper Chesapeake Bay area, aerial photos showed that common reed spread over 33 feet (10 m) in a single season on a bare sandy dredge [186]. Common reed clones on the Connecticut River's east shore from Long Island Sound to Lord's Cove spread from 33 m² to 1,630 m²/year [240]. Rates of common reed spatial expansion were 0.07 to 1.3 feet (0.02-0.4 m)/year and perimeter expansion rates were 1.6 to 6.6 feet (0.5-2 m)/year in New Jersey and Delaware photos taken from 1954 to 2000 [180].

Vegetative spread allowed common reed to occupy harsh sites with salinity of 20,000 to 30,000 ppm and daily flooding. Researchers conducted transplant and rhizome severing studies in low (daily tide flooding) and high marshes (no daily flooding) in the brackish (<15,000 ppm) Adolf Rotundo Wildlife Sanctuary of Massachusetts and in the saltier (20,000-30,000 ppm) Rhode Island's Rumstick Cove. The density of transplant shoots in high marshes was 2 to 5 times that in low marshes with highly anoxic soils. At Rumstick Cove, severing common reed rhizomes decreased the survival (<45%) of ramets growing into the low marsh. In the Adolf Rotundo Sanctuary, rhizome severing did not affect survival. Higher salinity at Rumstick Cove likely made the connection to the parent plant more important for survival [5].

In the Riverbend Marsh of New Jersey's Hackensack Meadowlands, common reed's spread from mosquito ditches into high marshes was facilitated through its alteration of the site. Severing rhizomes and clipping dead culms led to increased sulfide concentrations in dense common reed stands. Researchers suggested that common reed plants lowered sulfide concentrations in the upper marsh surface through oxygenation and perhaps pressurized ventilation of the rhizosphere. Decreased sulfide levels were associated with increased common reed growth. Establishment occurred in well-drained mosquito ditches low in free sulfides, and established plants provided a source of essential nutrients to the advancing plants through their rhizome connection [19].

Growth: Common reed is capable of rapid above- and belowground growth, with growth rates of up to 1.6 inches (4 cm)/day reported [202]. Rapid common reed growth may affect nutrient availability. In the Tivoli Bays of the Hudson River National Estuarine Research Reserve in New York, common reed produced nearly twice the aboveground biomass of narrow-leaved cattail and purple loosestrife (Lythrum salicaria) and sequestered a significantly greater amount of nitrogen and phosphorus in aboveground tissue (P=0.0001) [220].

Native and nonnative plant growth: The nonnative common reed haplotype emerges earlier, produces greater biomass, and activates dormant rhizome buds more rapidly than native common reed haplotypes. Native haplotypes may also be more susceptible to aphid herbivory than the nonnative type.

Field and greenhouse studies of native and nonnative common reed populations growing together in the Appoquinimick River watershed near Odessa, Delaware, showed that nonnative plants emerged earlier, accumulated more biomass, grew taller, and activated dormant rhizome buds more rapidly than native plants. In March, nonnative stands averaged 97.5 aerial shoots/m² whereas native stands averaged 7.5 shoots/m². Nonnative plants emerged earlier and flowered later than native plants. Differences in stand densities were not detected at the end of the growing season, but in August, height, fresh biomass, leaf biomass, and stem biomass were significantly greater in nonnative than native stands (P<0.0001). After 70 days in a greenhouse, rhizomes collected from the nonnative stands had produced significantly more shoots/biomass of rhizome planted than rhizomes collected from native stands (P=0.024). Researchers concluded that nonnative plants activated dormant rhizome buds more rapidly than native plants [144].

Greenhouse and field studies revealed that aphids (Hyalopterus pruni) preferred to feed on and had greater densities in native than nonnative common reed stands in Rhode Island. In the greenhouse, there were significantly more aphids/gram of dry weight on native than nonnative plants (P=0.037). Aphid feeding led to chlorosis and sometimes death of native plants, while nonnative plants were "relatively undamaged". In the field, nonnative stands supported a significantly lower density of aphids than did native stands (P<0.001). The only plants without aphids were nonnative [139].

Throughout its range, common reed is most common on wet, muddy, or flooded areas around ponds, marshes, lakes, springs, irrigation ditches, and other waterways. Common reed tolerates brackish and saline conditions [15,51,63,112,181,190]. In a review, authors report that common reed grows best in areas with slow or stagnant water and silty substrates [114]. However, on the Delmarva Peninsula along the Atlantic Coast, native common reed populations were more common along rivers than in marshes [161].

Established clones typically tolerate harsher conditions than seedlings. A review reported that growth from established clones was much less restricted than that of seedlings or sprouts. Newly established plants were limited to sites with less than 10,000 ppm salinity, sulfide concentrations below 0.1 mM, and a flooding frequency of less than 10%. Established clones grew in salinity up to 45,000 ppm, sulfide concentrations above 1.75 mM, and continuous flooding [42].

Climate: The large range occupied by common reed implies a wide climatic tolerance. In North America, common reed occurs in semiarid to arid desert, subhumid to humid continental, and subtropical climates. References consulted throughout this review showed that climates in common reed habitats varied widely by region. Information on temperature ranges, annual precipitation, growing season length, and possible disturbance weather given in this literature are presented below. Minimum and maximum temperatures and precipitation levels reported are specific to the location identified and based on a finite time period.

Northern United States: Common reed habitats in the northern Great Lakes states experience a subhumid, continental climate. Summers are short and warm; winters are long and cold. Annual precipitation averages 20 inches (508 mm) in northwestern Minnesota and 33.9 inches (860 mm) in Michigan's upper peninsula. Most of the precipitation (66%) occurs from April to September [29]. In the Lake Agassiz Peatlands Natural Area of Minnesota, January minimum temperatures average -39 °F (-39 °C), and July maximums average 94 °F (34 °C) (review by [108]).

Great Basin and Mojave deserts: In Utah and Oregon, common reed can occupy habitats in arid and semiarid climates [176,207,248]. At Diamond Pond in Harney County, Oregon, relative humidity is low, evaporation is high, and the growing season is short (80-117 days). Annual rainfall averages 7.9 to 12 inches (200-300 mm). Daily and seasonal temperatures fluctuate widely [248]. In Death Valley, common reed grows, when water is abundant, in locations where July temperatures can reach 110 to 115 °F (43-46 °C) [176].

Southern United States: Common reed habitats in South Carolina experience a subtropical climate with long, hot, humid summers and mild winters. The growing season averages 254 days. Average annual precipitation is 49 inches (1,245 mm), and hurricanes are possible but infrequent [211]. Common reed is typical in coastal prairies along the Gulf Coast in southeastern Texas and Louisiana. The climate is subtropical humid to semiarid in the Gulf Coast. The frost-free period averages 240 days in Louisiana and more than 320 days in lower Texas. Annual precipitation averages 56.6 inches (1,437 mm) at Lake Charles, Louisiana, and 28.8 inches (732 mm) at Corpus Christi, Texas. Ice storms, tropical storms, and hurricanes are possible (review by [205]).

A few studies have focused on the effect of specific weather and climates on common reed survival and growth. Several years of observations and studies in England indicated that spring frosts often increased common reed shoot density, crop biomass, and emergence period but decreased stem height and diameter [101]. Common reed plants taken from a Nebraska Sandhills meadow rolled their leaves when subjected to drought stress. Leaf rolling decreased the leaf area exposed to radiation [96]. Common reed growth and reproduction were greatest during an El Niño year in southern New England. Growth and reproduction were compared for 3 years, beginning 1 year before a high-precipitation El Niño year. Spring and summer were dry in the year before the El Niño. In the El Niño year, winter and spring were among the 10 hottest and wettest in the past 105 years. The following year had the 3rd hottest and 8th driest conditions in a 105-year period. On average, 30% more shoots were produced, shoots were 25% taller, and 10 times as many inflorescences were produced in the El Niño year than in years before or after. Soil salinity was negatively related to precipitation over the 3 years, and decreased salinity through precipitation inputs may have improved common reed growth [164].

Elevation: Common reed occupies sites from sea level to 7,000 feet (2,100 m) throughout North America. Elevation ranges in specific geographical areas are given below.

Elevational range of common reed by state
State Elevation (feet)
California below 5,200 [59,111,169]
Colorado 3,500-6,500 [97]
Idaho (eastern) 3,200-5,280 [92]
Michigan below 4,900 [235]
Montana (central and eastern) 2,100-3,850 [95]
Nevada 2,000-6,700 [23,127]
New Mexico 3,500-6,000 [158]
Utah 2,500-6,500 [247]
Utah (Uinta Basin) below 7,000 [84]

Soils: Common reed occupies a wide variety of substrates and tolerates a range of nutrients, organic matter, and pH levels. Soils in common reed habitats are described as "tight" clays in north-central Texas [58], rich and moist in West Virginia [215], wet and moderately fertile in the Great Plains [216], peaty in salt marshes along the north Atlantic Coast [63], minerotrophic peats in the northern Great Lakes states [29], and seasonally flooded clay to sandy loams in southern and eastern Idaho and central and eastern Montana [92,94]. In temperate regions, common reed may form a floating mat or island that is not well rooted in the substrate [114,181].

Nutrients/pH: Soils in common reed habitats may be acidic, basic, nutrient rich or nutrient poor, but soil and water conditions tolerated may depend on developmental stage.

Stunted common reed plants grew on acid tailings from an abandoned copper mine in Vermont where the pH was 2.9 (Penko 1993, personal communication, cited in [155]). In Louisiana coastal marshes, common reed occupied sites with pH ranging from 3.7 to 8. Additional information on the soil nutrients in coastal marshes is available from Chabreck [39]. In the Fish Springs National Wildlife Refuge of Utah, common reed communities occurred where pH levels were 8.2 to 9.2 and organic matter was 4% to 4.6% [30]. In the Lake Agassiz Peatlands Natural Area of Minnesota, common reed was indicative of weakly minerotrophic waters with pH of 4.3 to 5.8 and calcium levels of 3 to 10 ppm [108]. In Wisconsin, common reed occurs in emergent aquatic communities in waters with less than 50 ppm and more than 150 ppm calcium carbonate [54]. Cover of common reed was significantly greater in undiked than diked wetlands on Lake Huron and Lake Michigan (P<0.0001). Diked wetlands had more stable water levels than undiked wetlands. Soils in diked wetlands were organic and in undiked wetlands were sandy or silty. Soils in diked areas were significantly more acidic, and had significantly more organic matter, total nitrogen, and available phosphorus than soils in undiked areas (P<0.001) [110].

Water level: Common reed tolerates frequent, prolonged flooding as well as seasonal drying [94,124]. The frequency, level, and duration of flooding tolerated by common reed differs by site. Flooding can also affect salinity levels. In the northeastern United States, common reed survival and growth were best at low salinity [37,109,134] and low flooding conditions. Growth was reduced by flooding at low salinity levels but increased with flooding at high salinity (>18,000 ppm) levels [37,134].

Common reed's tolerance of flooding frequency, level, and duration varies by site. Voss [235] reported that common reed occurred in water up to 6 feet deep in Michigan. Common reed occurred on sites with "frequent and prolonged" flooding in central and eastern Montana [94]. A review reported that common reed can survive flooding levels of 1 foot (0.3 m) or more for at least 8 years [156]. Southern cattail-common reed communities along the Colorado River in the Grand Canyon occur on sites that are inundated an average of 54% of the time [213]. Common reed plants collected from the Gulf of Mexico and grown in the greenhouse had greater average stem height when grown in 8 inches (20 cm) of water than plants kept moist (P<0.05) [116]. However, common reed was often killed when roots were submerged for repeated growing seasons in Manitoba's Delta Marsh [238], where it was most typical of moist sites and avoided areas with more than 1.6 feet (0.5 m) of summer water [150]. A review of prairie marshes of western Canada indicated that common reed did not persist where the water table was deeper than 39 inches (100 cm). Clones did not spread where the water table was more than 20 inches (50 cm) deep, and mortality was likely if plants were flooded for 3 years with more than 3 feet (1 m) of water [202].

Fluctuating water levels are also tolerated by common reed. In southern and eastern Idaho, the common reed habitat type occurs on seasonally flooded sites where water levels range from 20 inches (50 cm) above to 3 feet (1 m) below the soil surface [92]. Water levels in common reed habitats of the Rocky Mountain Region fluctuate from 2 feet (0.6 m) above to 2 feet (0.6 m) below the soil surface [124]. On the Tailhandier Flats on the St Lawrence River of Quebec, common reed persisted in dry (water table >3 feet (1 m) deep) and in flooded (8 inches (20 cm) deep for 90 days) conditions. Area occupied increased when low water levels occurred in the previous year's growing season and decreased when the water table was 4.9 feet (1.5 m) or more deep or when flooded for more than 100 growing-season days [117].

Salinity: While common reed tolerates high salinity levels (up to 45,000 ppm) [42], it typically grows and establishes best in sites with low salinity (0-5,000 ppm). Along Long Island Sound in Connecticut, common reed did not occur on sites with more than 26,000 ppm salinity. Common reed cover, frequency, stem height, and percentage of flowering stems were significantly negatively correlated with salinity (P0.003) [241]. In marshes along the Connecticut River, common reed was significantly taller and produced more biomass/ramet in fresh (0-5,000 ppm) than brackish (11,000-17,000 ppm) marshes (P<0.001). Shoots emerged significantly earlier in fresh than brackish marshes, but common reed stem density was significantly greater (P<0.0001) in brackish than freshwater [67]. On the Delmarva Peninsula, native common reed populations were most common in low salinity habitats [161]. In the upper Chesapeake Bay area, common reed colonized freshwater (0-200 ppm) before mesohaline (2,000-10,000 ppm) marshes based on aerial photos taken between 1938 and 1995 [186]. Common reed plants collected from the Gulf of Mexico and grown in the greenhouse in salinity of 4,000 or 10,000 ppm had lower total stem height than those grown without salt (P<0.05) [116].

Common reed is considered both a pioneer and a climax species. It regenerates and establishes well on disturbed sites and is often considered a weedy or nuisance species. Generally, common reed is shade intolerant, appears early in primary open water succession, and sprouts rapidly after top-killing disturbances.

General descriptions: In marsh successions, common reed may be present in any seral stage from pioneer to climax. In the Fish Springs National Wildlife Refuge of Juab County, Utah, common reed's presence could result from an invasion into any seral stage of marsh/meadow community development or could represent any seral stage in regular succession from pioneer to climax [30,53]. In south-central Manitoba's Delta Marshes, common reed regenerates rapidly after disturbances and is considered a climax species [238,239].

Several researchers and systematists have described common reed as "weedy" and "invasive" [111,184,230]. Common reed is often described as characteristic of disturbed sites [63,216,256]. These descriptions have been applied to both the native and nonnative common reed haplotypes. For a discussion on where the nonnative haplotype is most common, see Subspecies, variety, and haplotype distributions, and for differences between native and nonnative haplotype growth, see Native and nonnative seedling growth and Native and nonnative plant growth.

Shade tolerance: Common reed is most common in full sun or nearly full sun conditions [111]. A review reports that common reed height and density are lower in partial shade [133]. In the Crystal Fen of north-central Maine, common reed occurred in open and recently forested but not in long forested portions of the fen [120]. In England, common reed occurred in closed-canopy woodlands, but plants were "spare, short, and flaccid" [102].

Primary succession: Common reed is often present early in freshwater swamp succession in the Great Lakes area but may appear a bit later in salt marsh succession along the Atlantic and Gulf coasts.

Vegetation development on open water in deep swamps, lakes, ponds, swales, and marshes is typically initiated with the establishment of submerged leaf species such as watermilfoil and/or bladderwort (Myriophyllum and Utricularia spp.) and closely followed by the establishment of floating leaf species including waterlily, buttercup, and/or pondweed (Nymphaea, Ranunculus, and Potamogeton spp.). Common reed typically establishes after the floating leaf stage. Eventually swamps may succeed to meadows or deciduous forest. This type of hydrosere succession is common in the Great Lakes area [66,75].

Four successional stages are recognized in salt marsh succession along the Atlantic and Gulf coasts, and common reed occurs in later stages that dominate as salinity and flooding decrease. The earliest successional stage is dominated by smooth cordgrass (Spartina alterniflora) and experiences saltwater inundation for 20 hours/day. The 2nd stage is dominated by saltgrass where salinity ranges from 30,000 to 46,000 ppm, and the water table fluctuates 2 inches (5 cm) above or below the soil surface. Saltmeadow cordgrass dominates the 3rd stage, when salinities are 7,500 to 35,000 ppm, and water levels are 4 inches (10 cm) above or below the soil surface. Common reed is not typically present until the final stage of succession, when salinity levels drop to less than 21,000 ppm, and water levels are between 4 to 8 inches (10-20 cm) below and 2 to 3 inches (5-8 cm) above the soil surface. The 3rd and 4th stages of salt marsh succession are considered edaphic climaxes. Sites may succeed to shrubs and eventually to deciduous forest on the Atlantic Coast, but on the Gulf Coast, true prairie is the theoretical climax [4].

Secondary succession: Disturbed sites are often habitat for common reed. If dominant before a top-killing disturbance, common reed rapidly sprouts from surviving underground rhizomes and dominates again. If absent before a disturbance and a propagule source exists, common reed often establishes on disturbed and temporarily bare sites. In the Adolph Rotondo Wildlife Reserve along the Palmer River in Massachusetts, wrack (mats primarily composed of vegetation litter) stranded in marsh turf suppressed common reed growth but once the wrack was removed, bare spots were rapidly colonized by common reed [163].

Natural disturbances: Grazing, fires, storms, and scouring are common disturbances in common reed habitats and often reduce the density and cover of common reed for a short time. Multiple disturbances or long-duration disturbances may produce longer-lived or more substantial decreases in common reed density and/or cover.

In Montana and Idaho, young common reed stems are palatable to both livestock and wildlife, and heavy grazing may decrease the size and extent of stands [92,94]. In the southern United States, grazing deferments of 60 to 90 days every 2 to 3 years are recommended if managing to keep common reed stands [145]. In the Ottawa National Wildlife Refuge east of Toledo, Ohio, common reed cover was reduced by grazing and sediment disturbances. Grazing and soil disturbances were evaluated through the use of exclosures on the mudflat study site that eliminated goose and white-tailed deer grazing and by turning over the top 6 inches (15 cm) of soil to mimic the effects of a storm or floating ice sheet. Common reed cover was 0.2% in disturbed and grazed plots, 0.2% in disturbed and ungrazed plots, 0.8% in undisturbed, grazed plots, and 10% in undisturbed, ungrazed plots. Researchers indicated that grazing and sediment disturbances produced additive effects and significantly decreased common reed cover (P<0.001) [17]. In the Sandhills of Nebraska, common reed was important in only the least disturbed "relatively high quality" fen. Common reed did not occur in fens that had large areas mowed, hayed, and/or grazed or in a heavily cattle grazed fen that had been planted with nonnative species [32].

Fire is typically only a top-killing disturbance in common reed stands. New sprouts may appear in as few as 5 days after fire [238]. From studies and observations in England, Haslam [103] found that burning broke rhizome bud dormancy but that cutting had "little effect" on the internal dormancy of rhizomes. In bog forest succession in northern Minnesota, narrow-leaved cattail-common reed communities may replace forest vegetation when sites are flooded or when fires burn deep into the peat layer and water collects [210]. For a complete summary of common reed's response to fire, see Fire Effects.

Coastal storms often provide opportunities for common reed establishment and/or spread. On Wallops Island, Accomack County, Virginia, common reed rapidly colonized bare areas of sand deposited in a January storm. Common reed produced stolons up to 70 feet (20 m) long, and while some plants appeared "stressed", others had produced small patches of "healthy-looking" stems [2]. On the Virginia Coast Reserve, areas disturbed by thick mats of wrack washed up by storms or high water events are often colonized by common reed. It is possible that common reed rhizome pieces or seeds were present in the wrack (Truitt 1992, personal communication, cited in [155]). Hurricane Camille produced a short-lived decrease in common reed dominance along the Mississippi River Delta. Relative abundance of common reed "declined considerably after the hurricane; however, 1 year following the storm this plant showed practically no change in abundance." Water and soil salinity were higher for a short time after the hurricane [40].

On mid-Atlantic coast sites, new common reed patches were common within 20 feet (5 m) of creeks or drainages. While there was a high concentration of establishment along creek banks, spread was not concentrated along creek edges. Creek edges likely received heavy propagule dispersal pressure, but were not suitable for recruitment [143].

Anthropogenic disturbances: Common reed is often found on sites disturbed by human activities. Common reed was present on 24-year-old peat mine sites but was not present on 1-, 6-, or 10-year-old mine sites in Wainfleet Bog, southern Ontario. Sites had been cleared of all living vegetation, and peat up to 7 feet (2 m) deep was removed. Mined sites were left to regenerate naturally [125]. On Wallops Island, Accomack County, Virginia, Ailes [2] observed a "sharp rise in the extent" of common reed stands on areas bulldozed as a fire break. In wetlands of the Chesapeake Bay subestuaries, common reed abundance was substantially greater in developed areas than undeveloped areas [130].

Increases in the nonnative common reed haplotype have been related to increases in road, waterway, and housing construction. In salt marshes in Narragansett Bay, Rhode Island, there was a significant positive correlation between percentage of marsh perimeter from which woody vegetation had been removed and percentage of border dominated by the nonnative common reed haplotype (=0.9173, P<0.01). Removal of woody vegetation and development in the area typically decreased soil salinity and increased available nitrogen [204]. In Quebec, the nonnative common reed haplotype occurred in 1916 but was rare and restricted to shores of the St Lawrence River before the 1970s. Before 1950, 92% of common reed populations sampled were native. In early 2000, more than 95% of colonies sampled were nonnative. The nonnative haplotype was especially common along roads but also occurred in marshes. Researchers suggested that the nonnative increase was facilitated by times of low water in the St Lawrence River and draining, dredging, excavation, and landfill operations associated with agriculture, housing, and road construction [146].

Eleven of 15 constructed tidal wetlands in Virginia's Coastal Plain were colonized by common reed. Dramatic increases in common reed abundance typically did not occur until wetlands reached 6 years old. Wetlands (7-12 years old) with perimeter ditches had significantly less common reed than wetlands without perimeter ditches (P=0.046). Subtidal perimeter ditches may have restricted rhizome establishment and growth into interior wetlands [106]. When these wetlands were 10 to 15 years old, common reed had established or spread into another of the constructed wetlands, but abundance was lower on sites where common reed had been replaced by red maple (Acer rubrum) scrub [105].

Flowering and fruit development occur from July to November throughout common reed's range [58,62,82,169]. In Florida, common reed flowering may occur as early as May and as late as January. However, these dates were associated with range extremes and/or abnormal weather events or patterns [91].

Typical common reed flowering dates by state and region
State/region Flowering dates
Arizona July-October [128]
Baja California July-November [249]
California July-November [169]
Florida October-November in the panhandle [46];
fall throughout [256]
Illinois July-September [168]
Nevada July-November [127];
September-October at Nevada Test Site [23]
New Mexico August-October [158]
North and South Carolina September-October [182]
Texas July-November [58]
Utah* (Uinta Basin) July-October [84]
West Virginia July-September [215]
Atlantic and Gulf coasts July-October [63]
Great Plains July-September (June-October, occasionally) [87,142]
Intermountain West July-September [51]
New England August-November [82]
Northeast August-September; often flowers persist through winter [153]
*In western Utah, common reed growth begins 14 April in a normal year and 2 weeks earlier in warm year; anthesis begins 15 July in warm year [30]


SPECIES: Phragmites australis
Fire adaptations: After fire in established common reed stands, new stems normally sprout from surviving rhizomes. Rhizome damage from deep burning may reduce common reed density and/or increase recovery time; however, lethal temperatures penetrating deep into the soil are rare in wet to moist common reed habitats [88,207,208,238]. New establishment on burned sites is possible given a viable seed or rhizome source. For more information on common reed establishment from seeds or rhizomes, see Regeneration Processes. Additional information about common reed's response to fire is available in Fire Effects.

Fire regimes: Fuels in common reed stands are conducive to flammability and fire spread. The high productivity and density of common reed stands provide fuel loads that are often higher than those of neighboring upland vegetation. In the upper Midwest, wetland fires can burn "hotter" and, given proper conditions, "faster" than fires in upland sites [188]. Common reed vegetation on the barrier islands of the Mid-Atlantic Coastal Plain is considered "extremely flammable" in the winter and early spring [83]. On unburned sites in the Delta Marsh of south-central Manitoba, common reed litter can be 18 inches (46 cm) deep [239]. On Cape Hatteras National Seashore, researchers indicated that fire carried even in flooded conditions provided dry litter was present [31].

©Gary Fewless
Cofrin Center for Biodiversity
University of Wisconsin-Green Bay

Pre- and early-settlement fires: Several studies report that Native people as well as early trappers and settlers burned wetland vegetation to improve travel, hunting success, and food availability.

California and Mexico: Native tribes of California burned common reed stands [8]. Rural people of Jaumave, Sierra Madre Oriental, Mexico, burned common reed stands to recycle nutrients, activate rhizomes, and reduce insect pests. Common reed sprouts were used as roofing and construction material [7].

Central Canada: In south-central Manitoba, Delta Marshes were intentionally burned by early trappers to improve travel, expose common muskrat lodges and coyote, fox, and American mink dens, and concentrate wildlife into unburned areas. Early settlers often burned Manitoba meadows to improve forage quality. Meadow fires often escaped and burned adjacent marshes. Burning was usually conducted in the first warm days of spring. Spring fires maintained common reed cover since they restricted the growth of encroaching woody vegetation and rarely killed belowground structures. Summer fires created temporary openings in common reed stands when they burned into peat and damaged rhizomes [238].

Southeast: Trappers burned marshes in southeastern Louisiana to improve trap accessibility and encourage growth of preferred common muskrat foods such as common reed. Fires typically burned when soils were wet and caused only minimal damage to marsh vegetation. Fires set after an extended drought, when peat and/or humus layers were dry, burned "furiously" [178]. In the southeastern United States, presettlement fire frequencies in brackish (5,000-30,000 ppm) and oligohaline (300-5,000 ppm) marshes that are typical common reed habitat ranged from 7 to more than 300 years; but fire intervals longer than 100 years were rare, and nearly all wetland sites including some islands had evidence of past fire. Fire frequency was estimated through a synthesis of information on soils, salinity, landscapes, remnant vegetation, historical records, and fire behavior in adjacent upland vegetation. Fires may have originated from burning in upland sites, lightning strikes, ignitions by Native Americans, or spontaneous combustion [76,77].

Spontaneous combustion was reported in marshlands along the shore of Lake Pontchartrain near Mandeville, Louisiana. Witnesses watched a fire "apparently ignited spontaneously" on 4 August 1924 in a time of "unprecedented drought". Water levels were several feet below the soil surface, and temperatures in neighboring towns were 100 to 104 °F (38-40 °C). Additional investigations in the area revealed that at least 100 separate fires were burning along an 18-mile stretch of marsh and pine vegetation. Other possible ignition sources were ruled out due to accessibility and timing constraints. Weather reports indicated that heating and ignition conditions necessary for spontaneous agricultural fires occurred that day near Lake Pontchartrain. Other naturalists in the area suggested that ignition may have come from a creeping ground fire [234].

Northeast: In New England and possibly other areas, proximity to a railroad may have increased fire frequency in common reed stands. Paleoecological studies in the Crystal Fen of north-central Maine showed that fire frequency increased after the construction of a railroad in 1893, then decreased sharply as spark-throwing steam engines were replaced by diesel engines [120]. In Massachusetts, 25% of all forest fires between 1916 and 1920 reportedly resulted from train engine ignitions (Averill and Frost 1933, cited in [120]).

Recent fire regimes: There is little information on current fire regimes in common reed habitats. Where common reed has spread into previously unoccupied areas, fuel characteristics may have changed and may contribute to changes in fire regimes. However, as of this writing (2008) these changes were not documented in the literature. On the southwestern portion of Long Island, New York, common reed and northern bayberry dominate Floyd Bennett Field. Portions of the Field burn each year in accidental human-caused fires. Common reed will probably replace northern bayberry, which does not recover as rapidly as common reed after fire [189]. From 1993 to 1998, there were 0 to 6 fires/year in the Rockefeller State Wildlife Refuge on the Gulf Coast Chenier Plain in southwestern Louisiana. Common reed cover is typically less than 10% in this area (Hess unpublished data, cited in [78]).

The following table provides fire regime information that may be relevant to common reed. Communities included in the table are those where common reed has the greatest potential as a persistent species. Fire regimes typical of common reed stands may be closely related to fire regimes in adjacent upland communities. Find further fire regime information for the plant communities in which this species may occur by entering the species name in the FEIS home page under "Find Fire Regimes".

Fire regime information on vegetation communities in which common reed may occur. For each community, fire regime characteristics are taken from the LANDFIRE Rapid Assessment Vegetation Models [141]. These vegetation models were developed by local experts using available literature, local data, and/or expert opinion as documented in the PDF file linked from the name of each Potential Natural Vegetation Group listed below. Cells are blank where information is not available in the Rapid Assessment Vegetation Model.
Pacific Northwest California Southwest Great Basin Northern Rockies
Northern Great Plains Great Lakes Northeast South-central US Southern Appalachians
Pacific Northwest
Vegetation Community (Potential Natural Vegetation Group) Fire severity* Fire regime characteristics
Percent of fires Mean interval
Minimum interval
Maximum interval
Northwest Grassland
Marsh Replacement 74% 7    
Mixed 26% 20    
Alpine and subalpine meadows and grasslands Replacement 68% 350 200 500
Mixed 32% 750 500 >1,000
Vegetation Community (Potential Natural Vegetation Group) Fire severity* Fire regime characteristics
Percent of fires Mean interval
Minimum interval
Maximum interval
California Grassland
Herbaceous wetland Replacement 70% 15    
Mixed 30% 35    
Wet mountain meadow-Lodgepole pine (subalpine) Replacement 21% 100    
Mixed 10% 200    
Surface or low 69% 30    
Vegetation Community (Potential Natural Vegetation Group) Fire severity* Fire regime characteristics
Percent of fires Mean interval
Minimum interval
Maximum interval
Southwest Forested
Riparian forest with conifers Replacement 100% 435 300 550
Riparian deciduous woodland Replacement 50% 110 15 200
Mixed 20% 275 25  
Surface or low 30% 180 10  
Great Basin
Vegetation Community (Potential Natural Vegetation Group) Fire severity* Fire regime characteristics
Percent of fires Mean interval
Minimum interval
Maximum interval
Great Basin Grassland
Great Basin grassland Replacement 33% 75 40 110
Mixed 67% 37 20 54
Mountain meadow (mesic to dry) Replacement 66% 31 15 45
Mixed 34% 59 30 90
Northern Rockies
Vegetation Community (Potential Natural Vegetation Group) Fire severity* Fire regime characteristics
Percent of fires Mean interval
Minimum interval
Maximum interval
Northern Rockies Shrubland
Riparian (Wyoming)
Mixed 100% 100 25 500
Northern Great Plains
Vegetation Community (Potential Natural Vegetation Group) Fire severity* Fire regime characteristics
Percent of fires Mean interval
Minimum interval
Maximum interval
Northern Plains Woodland
Northern Great Plains wooded draws and ravines Replacement 38% 45 30 100
Mixed 18% 94    
Surface or low 43% 40 10  
Great Plains floodplain Replacement 100% 500    
Great Lakes
Vegetation Community (Potential Natural Vegetation Group) Fire severity* Fire regime characteristics
Percent of fires Mean interval
Minimum interval
Maximum interval
Great Lakes Forested
Great Lakes floodplain forest
Mixed 7% 833    
Surface or low 93% 61    
Vegetation Community (Potential Natural Vegetation Group) Fire severity* Fire regime characteristics
Percent of fires Mean interval
Minimum interval
Maximum interval
Northeast Grassland
Northern coastal marsh Replacement 97% 7 2 50
Mixed 3% 265 20  
South-central US
Vegetation Community (Potential Natural Vegetation Group) Fire severity* Fire regime characteristics
Percent of fires Mean interval
Minimum interval
Maximum interval
South-central US Forested
Southern floodplain Replacement 42% 140    
Surface or low 58% 100    
Southern Appalachians
Vegetation Community (Potential Natural Vegetation Group) Fire severity* Fire regime characteristics
Percent of fires Mean interval
Minimum interval
Maximum interval
Southern Appalachians Forested
Bottomland hardwood forest Replacement 25% 435 200 >1,000
Mixed 24% 455 150 500
Surface or low 51% 210 50 250
Mixed mesophytic hardwood Replacement 11% 665    
Mixed 10% 715    
Surface or low 79% 90    
Vegetation Community (Potential Natural Vegetation Group) Fire severity* Fire regime characteristics
Percent of fires Mean interval
Minimum interval
Maximum interval
Southeast Grassland
Southeast Gulf Coastal Plain Blackland prairie and woodland Replacement 22% 7    
Mixed 78% 2.2    
Everglades sawgrass Replacement 96% 3 2 15
Surface or low 4% 70    
Floodplain marsh Replacement 100% 4 3 30
Pond cypress savanna Replacement 17% 120    
Mixed 27% 75    
Surface or low 57% 35    
Southern tidal brackish to freshwater marsh Replacement 100% 5    
Gulf Coast wet pine savanna Replacement 2% 165 10 500
Mixed 1% 500    
Surface or low 98% 3 1 10
Southeast Shrubland
Pocosin Replacement 1% >1,000 30 >1,000
Mixed 99% 12 3 20
Southeast Woodland
Atlantic wet pine savanna Replacement 4% 100    
Mixed 2% 175    
Surface or low 94% 4     
Southeast Forested
Maritime forest Replacement 18% 40   500
Mixed 2% 310 100 500
Surface or low 80% 9 3 50
Southern floodplain Replacement 7% 900    
Surface or low 93% 63    
*Fire Severities:
Replacement=Any fire that causes greater than 75% top removal of a vegetation-fuel type, resulting in general replacement of existing vegetation; may or may not cause a lethal effect on the plants.
Mixed=Any fire burning more than 5% of an area that does not qualify as a replacement, surface, or low-severity fire; includes mosaic and other fires that are intermediate in effects
Surface or low=Any fire that causes less than 25% upper layer replacement and/or removal in a vegetation-fuel class but burns 5% or more of the area. [93,140].

Rhizomatous herb, rhizome in soil
Ground residual colonizer (on site, initial community)
Secondary colonizer (on- or off-site seed sources)


SPECIES: Phragmites australis
Common reed is top-killed by fire, but rhizomes typically survive [103,135,238]. Although damage or death to common reed rhizomes is possible, it is not common.

Research from England indicates that burning common reed breaks rhizome internal dormancy [103]. Slight scorching by spring fires in Britain increased rhizome bud formation by as much as 400% (Haslam 1969, cited in [99]). Fires that burn deep into peat layers and/or burn during very dry conditions may damage or cause some mortality of common reed rhizomes [135].

Common reed sprouts rapidly from surviving rhizomes after fire. Sprouts may appear as soon as 5 days after fire [238]. Rarely is common reed abundance decreased by fire, and postfire recovery is typically rapid. At the end of the first season after fall and spring fires in the Delta Marsh of Manitoba, common reed shoots showed evidence of some scorching but survived to maturity. Fire-caused apical bud mortality was minimal [88]. If rhizomes are damaged or killed, common reed abundance may be reduced temporarily and/or recovery may be delayed [135], (review by [228]). Literature from northern mixed-grass prairies suggests summer fires (June-August) on dry substrates when plant nutrient reserves are low may burn into the organic soil and reduce common reed density through rhizome death or damage [135].

New common reed establishment on burned sites is possible if a viable seed or rhizome source exists. Seedling establishment is possible from on-site seed sources, but information on common reed seed banking is sparse. Establishment from rhizome fragments may be more successful than establishment from seed. Common reed plants established from rhizome pieces but not from seeds on burned soil in greenhouse and field studies conducted in Stemmers Run Wildlife Management Area, Maryland. Buried rhizomes had 100% survival in burned soils in the greenhouse. In the field, survival of sprouts from rhizomes on burned sites was 10%. Although no seedlings established on burned soils, 0.7% of seedlings established on bare mineral soil in the field [3]. For more information on common reed establishment from seeds or rhizomes, see Regeneration Processes.

The majority of fire studies indicate that common reed postfire abundance (cover, biomass production, and/or density) is rarely different from prefire abundance by the 2nd or 3rd postfire year [2,238]. It is not uncommon for burned sites to have greater common reed abundance than unburned sites [88,221]. Common reed abundance may decrease after summer fires, but decreases are likely short-lived; however, postfire measurements beyond the 3rd postfire year are rare [53,238]. This pattern is illustrated by reports from Utah, Manitoba, Virginia, North Carolina, and Delaware. These studies were conducted in a small portion of common reed's range. While the response to fire may be similar in other areas, additional studies are needed. There is also a lack of information of the effects of repeated fire in common reed habitats.

Utah: Common reed density increased after some summer and a late spring prescribed fires in the Fish Springs National Wildlife Refuge of Utah, but density was nearly 5 times lower in the first postfire year after the most severe summer fire. Fires killed over 90% of aboveground stems on all burned sites. Common reed density was greater in the first postfire year after fires on 15 June, 9 August, and 24 August and lower after fires on 29 June, 13 July, and 27 July. Decreased density after the 29 June, 13 July, and 27 July fires was not apparent at the end of the first postfire growing season, suggesting some delayed mortality. Common reed stem heights 1 year after fire were less than half of prefire heights on nearly all burns. On unburned plots, common reed density increased slightly from the prefire to the first postfire year, suggesting normal growing conditions. The prescribed fire of 13 July was the most severe and resulted in the largest decrease in common reed density. Peat soils smoldered for weeks and damage to rhizomes was noted. Below is a summary of common reed stem height and density on burned and unburned plots [53].

Prefire, postfire, and unburned common reed stem heights and densities [53]
Fire date (1981) Live stem height (cm) Density (live stems/m²)
Prefire End of first postfire growing season Following June Prefire End of first postfire growing season Following June
15 June 168 137 81 30 40 44
29 June 172 123 61 52 84 36
13 July 204 126 46 51 49 12
27 July 215 94 76 63 62 45
9 August 234 96 90 41 74 92
24 August 177 74 99 92 59 93
Control 1 246 247 138 60 50 64
Control 2 163 190 114 40 38 42
Control 3 207 218 144 40 34 47

Delta Marsh, Manitoba: Fire research in the Delta Marsh indicates that common reed density tends to be greater on burned than unburned sites regardless of fire season [88,221]. Decreases in common reed density were only reported after summer fires during dry conditions, and decreases were short-lived [238].

Common reed's shoot density and aboveground biomass were greater in the first postfire year after fall and spring fires than on unburned Delta Marsh sites, and common reed growth and reproductive development began earlier on burned than unburned sites. The fall fire removed almost all aboveground material, blackened the soil, but consumed little to no topsoil. Marshes were wet during the spring fire and there were unburned patches within the burned area. Fuels were not measured before either fire, but nearby unburned plots contained abundant dead plant material consisting of previous year's stems up to 7 feet (2 m) tall, and a litter layer almost 2 feet (0.5 m) deep [88]. Conditions during the fall and spring fires are provided below.

Conditions during fall and spring fires in the Delta Marsh [88]
  Fall fire
Spring fire
Air temperatures
(daytime high/nighttime low)
16 °C/-5 °C 13 °C/-5 °C
Wind speeds 5.5-19 km/h 5.5-19 km/h
Relative humidity 26-94% 27-100%

The first postfire growing season was warmer and 142% wetter than normal from April to August. The water table on fall-burned sites was 1 inch (2.5 cm) below the soil surface, on spring-burned sites was up to 15 inches (38 cm) below the surface, and on unburned sites was 20 inches (50 cm) below the surface. Common reed emerging after fire had some scorching but survived to maturity. Common reed density was greatest on fall-burned sites and averaged more than twice that of spring-burned and unburned sites. Aboveground biomass averaged 791 and 734 g/m² on fall-burned, 588 and 785 g/m² on spring-burned, and 402 to 423 g/m² on unburned sites. Average flowering shoot density was 20 to 30 stems/m² greater on fall-burned than unburned sites. Flowering was earliest on fall-burned sites, about a week later on spring-burned sites, and about 2 weeks later on unburned sites [88].

Common reed plants were slightly smaller but density was greater on burned than unburned sites in the Delta Marsh after spring, summer, and fall prescribed fires. Summer (1 August) and fall (7 October) fires burned in a year of periodic flooding, and the spring fire (11 May) burned in a year when sites were not flooded. Fires removed more than 90% of living and dead material in common reed stands and produced soil surface temperatures of 480 to 930 °F (250-500 °C). Postfire measurements were made about 3 to 4 months after the spring fire, 1 year after the summer fire, and 10 to 11 months after the fall fire. Common reed shoots emerged 19 May on spring-burned, 1 May on summer- and fall-burned, and 26 May on unburned sites. Regardless of burn season, common reed vegetative shoot density was at least 5 times greater on burned than unburned sites (P<0.05). The density, biomass, and proportion of flowering shoots were lower on summer- and fall-burned than unburned sites. These values were not different between spring-burned and unburned sites. Total aboveground common reed biomass was significantly greater on spring-burned, significantly lower on summer-burned, and not significantly different on fall-burned sites when compared to unburned sites (P<0.05). Community richness, evenness, and diversity increased on summer-burned plots [222]. Researchers thought that litter removal allowed for increased shoot density [221,222]. Information on these fires' effects on soil nutrients, soil temperatures, and aboveground plant nutrients is available from Thompson [223].

Average growth characteristics of unburned and burned common reed stands [221,222]
  Unburned Spring burned Fall burned Summer burned
Time since fire NA 3-4 months 10-11 months 1 year
Aboveground characteristics
Flowering shoots
Height (cm) 191.3a 179.0b 161.5c 141.1d
Basal diameter (mm) 6.6b 7.1a 6.7ab 5.7c
Inflorescence length (cm) 16.9a 17.1a 13.8b 12.9b
Leaf length (cm) 34.1a 34.1a 32.0b 31.6b
Leaf number 14.3a 13.6b 13.8ab 13.4b
Density (shoots/m²) 27.2ab 35.8a 16.0b 6.2c
Biomass (g/m²) 445.6a 431.9a 166.2b 49.1c
Vegetative shoots
Height (cm) 182.5a 151.3bc 163.0b 140.2c
Basal diameter (mm) 5.2bc 5.2bc 6.1a 5.0c
Density (shoots/m²) 17.5a 106.7b 105.8b 102.5b
Biomass (g/m²) 187.3a 482.6bc 575.2b 367.8c
Belowground characteristics
Standing crop (g/m²) 1,097b 1,880a 1,865a 1,208b
Bud density (buds/m²) 89b 206a 210a 115b
Different letters are significantly different (P<0.05) by fire treatment.

Two other summer fires in the Delta Marsh produced decreases in common reed density and stem height, and prefire density was regained by the 3rd postfire year after only 1 of the 2 fires. The 1st summer fire burned on 14 July when the air temperature was 90 °F (32 °C), and the wind speed was less than 15 miles (24 km)/h. In May, before the fire, the area to be burned was drained in an effort to make sites as dry as possible. Sprouts were visible within 5 days of the fire. By the 1st postfire frost, common reed stand height was half that of prefire height, and stem density was 66% less than prefire density. Prefire stand height and density were regained by the 3rd postfire year. The 2nd fire burned on 21 July in a drained area. Again common reed sprouts were present 5 days after the fire. Average stem height at the time of the first postfire frost was 22.7 inches (57.7 cm), and stem heights on a nearby unburned site were 84.2 inches (214 cm). The average number of common reed stems in a 22 × 14 inch (56 × 36 cm) frame was 29.6 on unburned sites and 5.7 in the 1st postfire year on burned sites. Common reed height and density were lower in burned than unburned sites in the 3rd postfire year [238]. These fires likely burned during much drier conditions than the summer fire described in the study above [221], which burned in a year of periodic flooding. Increased substrate dryness may explain the decreased common reed density after these summer fires. Also the researcher noted an increased use of burned areas by ducks and common muskrats after the 2 summer fires [238], whereas postfire grazing was not noted in the above study [221].

Mid-Atlantic Coast: Fall and spring fires did not significantly affect common reed abundance in Virginia and North Carolina. Relative common reed cover on burned and nearby similar unburned sites were not different 2 years after a 23 November prescribed fire in dune swale communities on Wallops Island, Virginia [2]. Common reed density was slightly lower, though not significantly, on burned than unburned plots after a late-April fire in common reed stands on Cape Hatteras National Seashore [31].

Comparisons of the establishment and spread of common reed populations were made from time series maps of the mid-Atlantic coast. At only 1 of 6 sites, Lang Tract, Delaware, did common reed abundance decrease for any length of time. Common reed was not present on a 1982 map likely due to a prescribed or wildfire. On a 1989 map, common reed was again dominant [143]. The fire was not described.

Repeated fire: Information on the effect of repeated fires in common reed habitats is lacking. Most landowners noted that common reed's introduction and spread on the salt hay farms in Commercial Township, New Jersey, coincided with Hurricane Hazel (1954); however, one individual noted that common reed's spread rate increased as fire frequency in the area decreased. It was not presumed that reduced fire frequency was the single or even primary reason for increased spread of common reed. Salt hay farms were, however, burned frequently in winter fires prior to the hurricane. The lack of information on the effects of repeated fire in common reed habitats makes it unclear whether or not repeated fire could decrease the spread of common reed [18].

Climate/weather: Research from the Netherlands and Britain indicate that climatic conditions during and after fires can affect the postfire development of common reed. In Britain, winter fires typically encourage early spring emergence, and spring fires can increase common reed stand density [99]. In Flevoland, the Netherlands, postfire frost damage was most severe on dry-burned sites. Greater decreases in standing dead material and litter on dry sites allowed for colder minimum temperatures than on wet-burned and unburned plots [226].

Common reed stands are not usually difficult to burn. Fuel loads are generally high, and only in recently burned sites does fire fail to spread. Additional information on fuel loadings in common reed stands is available in Fuels. Prescribed fires during very dry conditions or in conjunction with other control methods have been used successfully to reduce the size and/or spread of common reed stands. However, adverse impacts on wildlife are possible when burning common reed stands.

Conducting prescribed fire: Several challenges could make prescribed burning in common reed habitats difficult. High-intensity updrafts are possible in wetland habitats, and embers may move long distances [188]. Spot fires are possible 100 feet (30 m) from the burned area [228]. Firelines may need to be wider than those typically constructed in upper Midwest upland habitats. Maneuverability of water tanks can be compromised in wetlands and may increase the number of personnel needed to control fires in common reed habitats [188].

On Cape Hatteras National Seashore, prescribed fires burned in flooded conditions, and "wetline(s)" were constructed simply by trampling neighboring vegetation [31]. Although fires typically carry well in common reed habitats, there may be insufficient litter and dead material to burn in consecutive years. A 2nd winter fire was unsuccessful in the Nebraska Sandhills 1 year after a prescribed fire in common reed marsh due to sparse stems and a lack of accumulated litter. Common reed on the previously burned site "did not appear nearly as combustible as the old growth even when the flame was applied directly" [199].

The only study to report soil temperatures produced by prescribed fires in common reed habitats indicates that heat does not penetrate deeply. In a common reed stand in Utah's Ogden Bay Waterfowl Management Area, an early-September fire produced temperatures of 120 °F (48 °C) at 9.3 inches (23.7 cm) deep, 219 °F (104 °C) at 3 inches (7.7 cm) deep, 306 °F (152 °C) at 1.1 inch (3 cm) deep, and a high temperature of 399 °F (204 °C) penetrated only 0.2 inch (0.5 cm). The fire burned when wind speeds averaged 10.3 miles (16.6 km)/hour, the average dew point was 41 °F (5 °C), and the maximum daytime temperature was 83 °F (28.5 °C). Drawdown began in April on the burned sites, but canal leakage and precipitation were such that water pooled in pits [207].

Fire as a control method: Severe, deep-burning fires may kill common reed [208], and removal of thick common reed litter by fire may allow other species to establish [228]. In Atlantic Coast marshes, "root burns" and "peat fires" can be used to cause common reed rhizome mortality. "Root burns" require a completely dry marsh floor. "Peat fires" require several years of litter accumulation, a "fairly deep" peat layer, and drought conditions to sustain smoldering and deep burning [208].

In the early 1940s, spring and late-summer fires were used in the Delta Marsh to create open water sites, thin dense stands, and increase edge habitats, in order to benefit wildlife. Successful spring fires required a "stiff" wind and 2 to 3 days of warm, sunny weather to dry dead stems [237]. Spring fires during "dull days" often did not carry well and produced patchy burns [238]. With enough wind, fires would burn even when there was snow and/or water at the base of the plants. Spring fires did not usually damage common reed rhizomes and served to increase the proportion of edge habitat. Late-summer fires typically burned deep into the peat layer producing some rhizome mortality and creating open water in common reed stands. Successful summer fires required dry conditions, a dense stand, and sustained smoldering. Summer fires were typically set in late August or early September [237].

Fire in conjunction with other physical, mechanical, or chemical control methods may produce common reed mortality [3,18,31,155,171]. On Cape Hatteras National Seashore, repeated cutting of common reed on burned sites decreased its growth rate but did not cause mortality [31]. In the Stemmers Run Wildlife Management Area in Cecil County, Maryland, common reed abundance was reduced on sites that were burned 4 months after herbicide treatments. In the 4th posttreatment year, there were 275 common reed individuals in the total 58 quadrats (3.16 ×0.32 m) on treated sites. The number of individuals before treatments was 878 [3]. In oligohaline, wind-tide marshes in southeastern Virginia, common reed density and frequency were significantly reduced when sites were treated with a dormant-season fire between 2 herbicide treatments late in the growing season (P-value not reported). Herbicide treatments alone did not produce significant decreases from pretreatment levels [44].

Flooding burned sites can produce common reed mortality by eliminating oxygen transport from aboveground plant structures to roots and rhizomes [18]. "Snorkels are snipped" when burned sites are flooded (Gallagher, personal communication, cited in [18]). Several studies report this effect, though none provided details about fire or flooding conditions. In sawgrass-common reed vegetation in Louisiana coastal marshes, postfire flooding with saline water can produce mortality and reduce stand density [171]. In Connecticut, late-spring fires followed by saltwater flooding decreased the height and density of common reed stands (Steinki 1992, personal communication, cited in [155]). On the Wertheim National Wildlife Refuge in New York, common reed was eliminated for at least 3 years when portions of a freshwater impoundment were reflooded after winter burning that followed fall draining (Parris 1991, personal communication, cited in [155]).

Wildlife considerations: Fires in common reed marshes can be used to benefit wildlife, but can also negatively impact nesting birds. Prescribed fires should avoid destroying currently used nesting habitat. Studies conducted in the 1960s and 1970s in the Delta Marsh indicated that spring fires before 20 April typically missed the beginning of mallard and northern pintail nesting. Impacts on nesting birds can be minimized if summer fires are ignited after gadwall and blue-winged teal have left their nests [238]. Fall fires can decrease snow retention and affect spring run off levels, which may affect the value of winter and spring wildlife habitats [239].


SPECIES: Phragmites australis
Nutria, common muskrats, birds, and cattle feed on common reed. Song sparrows (Klockner 1985, personal communication, cited in [155]) and waterfowl eat seeds [133,216]. Black-capped chickadees and other bird species feed on scales (Caetococcus phragmitidis) that commonly occur in common reed leaf sheaths [133]. Nutria and common muskrats consume rhizomes and stems [133,216,254].

Cover value: Common reed provides shade, nesting, and cover habitat for mammals, waterfowl, song birds, and fishes. Native ungulates, waterfowl, other birds, and small mammals utilize common reed stands for cover. Waterfowl, pheasants, and rabbits use cover at the margin of common reed stands throughout its range [156]. In valley habitats of Nevada, common reed is considered "excellent" Gambel's quail cover [89]. In Idaho, common reed stands provide "excellent" big game thermal and hiding cover, and waterfowl utilize stands for nesting and hiding [92]. Common reed provides good feeding and thermal cover for many bird and small mammal species in Montana and is good thermal cover for mule deer and white-tailed deer [95]. In the Delta Marsh, white-tailed deer utilize common reed stands for escape cover [238]. More specific cover information is provided in the following subsections.

Livestock: Some report that common reed has little to no forage value [62,85], but Leithead and others [145] claim common reed is "readily eaten by cattle and horses" in the southern United States. Stubbendieck and others [216] also report that cattle and horses consumed common reed before it matured.

Small mammals: Common reed provides habitat for white-footed mice and habitat and food for nutria and common muskrats. The white-footed mouse, a habitat generalist, often occurs in common reed freshwater tidal marshes along the Hudson River of New York [160]. Common muskrats feed on common reed stems and use stems in nest construction [156]. Common reed may also provide emergency common muskrat cover on Gulf Coast marshes when lower marshes are swept away by storms or when other habitats are overpopulated [152]. Common reed is considered an important nutria food in Louisiana (Harris and Webert 1962, cited in [131]). In marshes of Dorchester County, Maryland, spring and fall nutria diets contained large amounts of common reed. Over a 3-year period, common reed made up 5.9% of nutria's annual diet, but made up 33.2% of May and 19% of October diets [254].

Birds: Common reed provides food as well as nesting, roosting, and hunting habitats to a wide variety of bird species. Some studies, however, indicate that dense, monotypic common reed stands support lower avian diversity than other wetland habitats.

Red-winged and yellow-headed blackbirds frequently use common reed habitats in central and eastern Montana [94]. Along the Colorado River from the Arizona-Nevada to the United States-Mexico borders, common reed stands supported the lowest avian densities and diversities of the marsh types studied. However, common reed marshes were utilized by wading birds in the spring and visiting insectivores throughout the year. In the spring, Yuma clapper rails also used common reed habitats [6].

Common reed is not considered an important food source for ducks, according to studies from Louisiana [41] and Georgia [123], but provides important nesting habitat. Stands with open water are typically preferred to thick dense stands. In the prairie pothole region of the northern United States and southern Canada, semipermanent and permanent marshes with large stands of common reed are important habitats for flightless, molting adult ducks [218,237]. Common reed stands also provided an important barrier for marsh inhabitants by limiting intrusions from grazing animals and humans [237].

Nesting habitat: Throughout its range, common reed is utilized as nesting cover and material. On the Bear River Migratory Bird Refuge on the northeastern edge of Utah's Great Salt Lake, snowy egrets and other herons used broken common reed stems as nest material [253]. In the Great Plains, red-winged blackbirds "preferentially" nested in common reed vegetation [216]. On southwestern Louisiana's Gulf Coast, red-winged blackbirds and boat-tailed grackles frequently nested in cattail and/or common reed stands [78].

On Pea Patch Island in New Castle County, Delaware, 10 wading bird species nested in common reed vegetation during a 7-year study. Snowy egrets, cattle egrets, little blue herons, and black-crowned night-herons as well as small numbers of tricolored herons, yellow-crowned night-herons, and green herons nested in common reed marshes and in upland sites. Cattle egrets produced larger clutches and had greater hatching success in common reed marshes than on upland sites, while the opposite was true for little blue herons. Common reed stands provided important nest material for wading birds and provided a protecting buffer from upland human and pet traffic [174].

On Utah's Bear River Migratory Bird Refuge, 3% of all duck nests (mallards, gadwalls, pintails, redhead, and cinnamon teal) were in common reed stands, although common reed occupied only an estimated 1% of the marsh area. The fate of duck eggs on the refuge is reported for species and vegetation type by Williams and Marshall [253]. Mallards used common reed more in developed than in undeveloped areas of Beach Haven West, New Jersey. Common reed was the primary nesting cover in developed lagoons [70].

Canada geese preferred bulrush, broad-leaved cattail, and common river grass over common reed cover types in Marshy Point, Manitoba, but the common reed cover type was preferred over other grasses and woodlands [48]. Nesting ducks in the Delta Marsh of southern Manitoba "heavily" used the edges of common reed stands. Mallards extensively used edge habitats where common reed met meadow vegetation, and redheads and lesser scaups used edges that met open water. Of 147 land-nesting duck nests, 31% occurred on the edges of common reed stands at the Delta Marsh Duck Station. Canopies created by the previous year's snow-weighted common reed stems and patches of common reed within meadow vegetation were favored nest sites. Flightless ducks often used open water areas within common reed vegetation [237].

Foraging/roosting habitat: Short-eared owls, barn swallows, chimney swifts, and red-tailed hawks utilize common reed habitats for roosting or foraging. On the lower Columbia River in Multnomah County, Oregon, short-eared owls roosted in old fields dominated by common reed and thistles [219]. Barn swallows and chimney swifts used common reed marshes along the Hudson River for perching and foraging [160]. In the Hackensack Meadowlands of New Jersey, short-eared owls used 2- to 3-foot (0.6-0.9 m) tall common reed stands for winter roosting [33], and red-tailed hawks hunted in common reed marshes [34].

Aquatic animals: Reviews report that common reed stands provide important shade, shelter, and food for fishes [114] and that common reed litter provides food for mollusks, other crustaceans, and aquatic insects [133]. There is additional information on the nonnative common reed haplotype and aquatic organisms in Impacts on fish and other aquatic organisms.

Palatability/nutritional value: Common reed is not rated as a high-value or high-palatability livestock or wildlife food unless plants are young. Immature plants are considered palatable in southern and eastern Idaho [92]. In Montana, common reed is considered a fair food source for pronghorn and a poor food source for mule deer, white-tailed deer, and elk. Palatability is rated fair for horses and cattle and poor for domestic sheep [95]. In the southern United States, common reed is described as a "high-quality, warm-season forage," although mature plants are considered tough and unpalatable [145].

Several studies report on the nutrients available in common reed plants. Trends in crude protein, phosphorus, and digestibility levels of common reed in south-central North Dakota from late spring to early summer are available from Kirby and others [132]. Percent ash, carbon, and nitrogen in live and dead aboveground common reed material is reported for plants from Blackbird Creek Marsh in New Castle County, Delaware, by Rowman and Daiber [193]. Levels of nitrogen and carbon in belowground common reed biomass along the Atlantic coast of Delaware are reported by Gallagher and Plumley [80].

Ease of establishment, rapid vegetative spread, and high tolerance of disturbance make common reed an understandable choice for rehabilitation. However, these same traits make common reed a nuisance or weedy species in some areas. In natural or wild areas, the use of native common reed haplotypes may be required or preferred. For more information on the potential impacts of the nonnative common reed haplotype, see Impacts.

Common reed seeds, rhizomes, and plants have been used in restoration [113,122,173]. The extensive common reed rhizome network is useful for bank stabilization [92]. In Lake Mead coves, common reed was planted to provide fish cover. Survival ranged from 0% to 56%. Plants did not survive on steep sites with rapidly dropping water levels [50]. Once phosphogypsum and clay slurries were deposited on open pit phosphate mines in Beaufort County, North Carolina, common reed colonized rapidly. However, establishment of nonriverine wet hardwood oaks and shrubs was less successful when common reed was present [9].

Native people ate common reed rhizomes and seeds. They also used the plant material to treat stomach, ear, and tooth pains, and to construct pipestems, arrows, mats, nets, and prayer sticks [62,127,128,242].

Common reed was utilized as a food source and as a medicine by Native Americans. Shoots were eaten raw or cooked. Flour was made from dried shoots and rhizomes [62,64]. Common reed rhizomes provided a year-round food source. Seeds were harvested and ground into a high fiber meal [62]. In southern California, the Kawaiisu harvested and utilized sugar crystals that collected on common reed stems [257]. Paiute people used common reed's sugary sap to treat lung ailments, and the Apache used common reed rhizomes to treat diarrhea, stomach troubles, earaches, and toothaches [62].

Common reed plant material was used to construct various items that made food gathering, warfare, travel, and relaxation easier or more comfortable. Native people used common reed in fences, roofs, and baskets [62]. Common reed was also used as insulation, fuel, fertilizer, and mulch. Six hundred-year-old cigarettes found in Red Bow Cliff Dwellings, Arizona, were constructed of common reed stems [181]. The Kawaiisu of southern California used common reed stems to make arrows, fire drills, and pipes [257]. The Cahuilla, also of southern California, used common reed stems to make flutes, splints, and arrow shafts. Common reed was also used as a thatch in house construction. The soft, silky fibers, which remained after stems were soaked and the outer tissue layer was removed, were twisted into a strong cordage used to make carrying nets and hammocks [22]. The Navajo used common reed to make bird snares and arrows [65]. The Seri of the southwestern United States bundled common reed stems to make "seagoing reed boats". Boys used mesquite (Prosopis spp.) spines attached to common reed stems to catch small fish and crabs [68]. The Navajo used common reed to make prayer sticks that they used during the Mountain Chant Ceremony [65].

Increases in the amount and coverage of nonnative common reed haplotypes since the mid-1900s have prompted many investigations into its potential allelopathy, method of establishment and spread, impacts on native plant and animal species, and susceptibility to control.

Allelopathy: The only study to date assessing allelopathy in common reed suggests its rhizomes do not exude allelopathic chemicals. Researchers found that germination of saltgrass and saltmarsh bulrush (Schoenoplectus robustus) was not affected by watering with common reed rhizome leachate [61].

Many studies have quantified and traced the spread of common reed in the Great Lakes and Atlantic Coast areas where the nonnative common reed haplotype has become dominant. Establishment, spread, and increased dominance of common reed are often associated with anthropogenic disturbances, including land development, tidal manipulation, and waterway construction. For more on the establishment and spread of common reed, see General Distribution and Occurrence and Regeration Processes.

Impacts: Numerous changes can occur when common reed replaces other vegetation. Common reed has been called an "ecosystem engineer" [212]. Plant diversity, soil properties, sedimentation rates, bird and fish habitat use, and food webs may be altered when marshes are converted to dense, monotypic common reed stands.

Impacts on plant diversity: The growth of large monotypic common reed stands may be associated with decreased plant diversity. Through field and greenhouse experiments, researchers concluded that common reed litter was the most important factor in the exclusion of other brackish tidal marsh species. Seeds of triangle orache (Atriplex prostrata) and seaside goldenrod (Solidago sempervirens) established and grew well in soils collected from sites dominated by common reed or rush (Juncus spp.) in the Adolph Rotundo Wildlife Preserve in Massachusetts. Total biomass of both species was greatest in common reed soils. In field experiments, establishment of these forbs decreased significantly (P<0.05) with common reed litter regardless of the presence of common reed shoots. Forb establishment increased with the removal of common reed litter and stems [166].

All measures of plant species diversity were lowest in a marsh with the greatest average standing crop of common reed (1,742 g/m²) in East Harbor State Park, Ohio. The researcher stressed cause-effect relationship was not established but suggested that long-term common reed persistence may have reduced seed bank species richness [244]. In the Kampoosa Bog of Stockbridge, Massachusetts, species richness and evenness were not different between fen plots with or without common reed. However, the cover of characteristic fen species, water sedge (Carex aquatilis) and sweetgale (Myrica gale), was significantly lower on plots with common reed (P<0.05) [187].

Impacts on sediment properties: Some studies indicate that common reed may alter soil properties, salinity levels, and topographic relief when it replaces previously dominant vegetation. Water salinity, depth to water table, and topographic relief were significantly lower in stands dominated by common reed than stands dominated by saltmeadow cordgrass and saltgrass in brackish tidal marshes on Hog Island in southern New Jersey (P<0.01). All 3 variables were also negatively correlated with common reed age. Significant differences in soil properties were noticed within 3 years of common reed establishment [255].

Stanton [212] described common reed as an "ecosystem engineer" after finding that true elevation, peat accumulation, and organic matter increased while sediment bulk density decreased with increased common reed dominance in southwestern Louisiana's Rockefeller Wildlife Refuge. Soils and elevation changes were compared along a gradient that included marshes dominated by saltmeadow cordgrass, saltgrass, and saltmarsh bulrush, ecotones between uninvaded marshes and marshes with new common reed establishment, and a monotypic common reed stand about 40 years old. Rates of elevation increase peaked within 7 years of common reed establishment. Sediment bulk density decreased with increased common reed age [212].

Common reed's impacts on sediment properties, however, are not consistently demonstrated over all studies and sites. In Maryland's Prospect Bay, flow regime, sediment transport, and sediment deposition patterns were not different at the scales measured in common reed and smooth cordgrass marshes. Researchers suggested that results may be different during severe storms [147]. In Tivoli North Bay, New York, there were no significant differences in sediment microbial biomass and activities among narrow-leaved cattail (Typha angustifolia), purple loosestrife, and common reed marshes. Microbial processes specific to pollutants were not studied and the study was conducted at the height of the growing season. Both factors may have affected findings [172].

Impacts on animal habitat: Conversion of wetland habitats to monotypic common reed stands may or may not affect animal use. Findings often differed with the species and age of the animal and vegetation being studied. In many cases, habitat diversity, size, and connectedness may affect wildlife more than plant species composition.

Birds and small mammals: In 40 salt and brackish marshes of Connecticut's tidal wetlands, there were significantly fewer state-listed (endangered, threatened, or special concern) bird species in common reed than in shortgrass vegetation dominated by saltmeadow rush, saltgrass, and/or cordgrass (P<0.001). The average number of bird species/plot was also significantly lower in common reed than shortgrass marshes (P=0.029). Bird communities in common reed vegetation were dominated by marsh wrens, red-winged blackbirds, swamp sparrows, and tree and barn swallows; wading birds and sandpipers foraged at the edge of common reed stands [24].

Along the Hudson River of New York, bird and small mammal species richness, species composition, and abundance were not significantly different between common reed, purple loosestrife, and cattail freshwater tidal marshes (P<0.05). Average bird species richness was highest in common reed marshes, although not significantly. Arthropod availability and nest predator access were also not different by vegetation type. Bird and arthropod abundance were better predicted by site and landscape characteristics than vegetation type [160].

Fish and other aquatic organisms: Habitat use by fish, crustaceans, and other aquatic invertebrates can be affected by vegetation; however, fish age as well as vegetation type may affect study findings. In a review, authors report that common reed marshes support a "diverse and abundant benthic biota", and that many estuarine organisms are not affected by common reed's presence [243]. On the East shore of the Connecticut River on Long Island Sound, common reed vegetation supported macroinvertebrate densities similar to those of restored meadows and smooth cordgrass-cattail vegetation [240]. On the Hog Islands of southern New Jersey, overall small fish (P=0.0001) and crustacean (P=0.002) use were significantly greater in smooth cordgrass than common reed vegetation [1]. Total fishes caught/trap was not significantly different between common reed and narrow-leaved cattail marshes (P<0.05); however, there were species-specific differences between the 2 vegetation types. The number of aquatic invertebrates collected per litter bag was generally highest in narrow-leaved cattail marshes, but differences between the 2 marsh types were not significant. Grass shrimp (Palaemonetes pugio) captures/trap were significantly greater in common reed than narrow-leaved cattail marshes (P=0.002). Fiddler crabs (Uca minax) were significantly more abundant in narrow-leaved cattail than common reed marshes (P<0.001) [69].

Several studies report that common reed-dominated marshes provide less suitable habitat for mummichog (Fundulus heteroclitus and F. luciae) larval and small juvenile forms [1,183]. Fundulus luciae was captured exclusively from smooth cordgrass marshes, and the abundance of recently hatched F. heteroclitus was much lower in common reed than smooth cordgrass [1]. Findings were similar along the Lieutenant River of Connecticut, where significantly more F. heteroclitus larvae and juveniles were caught from narrow-leaved cattail than common reed marshes (P<0.001) [69]. Successful pit trap of F. heteroclitus and F. luciae decreased with increased abundance of common reed in estuarine habitats in New Jersey, Delaware, and Maryland. Researchers suggested that increased litter accumulations in common reed marshes created a more uniform topography, decreased pooling, and may have reduced abundance of refugia from currents [118]. Along Mill Creek, in New Jersey's Hackensack Meadowlands, large juvenile and adult F. heteroclitus abundance was similar in common reed and smooth cordgrass marshes but larvae and small juveniles were significantly more abundant in smooth cordgrass than common reed (P=0.04 in 1999; P<0.0001 in 2000). Of 1,469 total fish captured, only 29 young of the year were captured from common reed marsh, and their most likely prey were significantly more abundant in smooth cordgrass than common reed (P<0.05). Experimentally creating undulations and pools in the sediment increased larval abundance some, but researchers cautioned that these findings do not indicate the undulations and pools are the only important larval habitat features [183].

Impacts on food webs: Arthropod food webs differed between smooth cordgrass and common reed stands in the Alloway Creek Watershed of New Jersey's Delaware Bay. In smooth cordgrass stands, the food web depended on herbivores and smooth cordgrass consumption. In common reed stands, a detritus-based food web was most common [86].

Control: While several studies report on the use of chemical, mechanical, and integrated control methods for common reed, determination of the common reed haplotype and assessment of potentially undesirable consequences of removal are necessary before control is attempted. In the Great Lakes area, on the Atlantic Coast, and in other parts of common reed's range, appropriate management of common reed requires that its native or nonnative status be determined. In some areas, land managers are attempting to maintain and encourage native common reed populations while discouraging nonnative populations [175].

Although common reed can be a problem in waterways, producing extensive stands that restrict water flow, the same aggressive growth characteristics make it an excellent soil binder that prevents erosion and washouts [114] and may protect eroding coastlines [191,192]. Therefore the control or removal of common reed may negatively impact some coastal locations. At eroding island sites on the eastern shore of Chesapeake Bay, Maryland, more deposition occurred in common reed than cordgrass stands. Common reed stands trapped minerals and organic sediments at a rate of 24 g/m²/day. Substrate elevation increased by as much as 3 mm in 6 months in common reed stands [191]. Additional studies in Chesapeake Bay showed that accretion rates were higher (0.95 cm/year) and sediment water content lower (about 70%) in 20-year-old common reed than in cattail, switchgrass, or 5-year-old common reed stands. High productivity, litter accumulations, and high sediment loadings in common reed marshes likely contributed to accretion. Researchers indicated that high accretion in common reed stands may actually benefit coastal areas since sea level rise in Chesapeake Bay is 2 to 3 times the eustatic rate of 1 to 2 mm/year [192], (sea level data reviewed in [192]).

Best management practices in common reed marshes may not require vegetation type conversions. In Delaware Bay estuaries and Connecticut River salt marshes, researchers assessed habitat data from common reed stands with intermittent and continuous herbicide use. Habitat value was rarely 0% or 100%, regardless of species composition and dominance, and smooth cordgrass did not colonize sprayed common reed zones as rapidly as cover was lost to herbicide treatment. Researchers suggested managing for a net gain of suitable habitat instead of a vegetation type conversion in these marshes [229]. In a review, Ludwig and others [151] suggested that common reed management should be site-specific, goal-specific, and value-driven. Understanding the biological, chemical, and physical impacts of common reed at a particular site is important to the management decision-making process [151].

Numerous studies have assessed control methods for common reed. Information on many individual and integrated methods is available from the following references: [52,155,227]. Some indicate that control treatments are most effective when plants are releasing pollen, typically in midsummer [156], and that extensive and persistent rhizomes necessitate follow-up treatments [57].

Prevention: Maintaining competing vegetation around existing common reed stands and minimizing nutrient loads may limit common reed spread. In a coastal brackish marsh along the Barrington River in Seekonk, Rhode Island, cutting neighboring vegetation and adding nutrients increased common reed (likely the nonnative haplotype) density, height, and biomass. Common reed spread 3 times farther in high-nutrient vegetation-removal treatments than in any nutrient treatment with intact neighboring vegetation [165].

Water level manipulation: In some areas of Connecticut, the reintroduction of tidal flooding through dike breaching has decreased the area occupied by common reed [241]. However, it is suggested that restoring fluctuating water levels in Great Lakes wetlands may increase common reed abundance [110].

Along Long Island Sound in Connecticut, breaching dikes that were more than 50 years old generally decreased the total marsh area covered by common reed. Through tide restoration, salt marsh vegetation replaced common reed at a rate of 0.5% to 5% per year and limited common reed to less frequently flooded sites [241]. In the Barn Island tidal marsh complex of Stonington, Connecticut, the reintroduction of tidal flooding decreased common reed abundance in places. Before dike construction, stunted smooth cordgrass, saltmeadow cordgrass, and saltgrass dominated. Thirty years after dike construction, cattail and common reed dominated. Ten years after tidal flooding was restored, 28% of the study area resembled predike vegetation, and 33% remained dominated by cattail and common reed [16].

Integrated management: Many studies describe the effects of multiple control methods on common reed. On Connecticut River's east shore, mowing and herbicide treatments provided for short-term control [240]. In common reed marshes near Salem, New Jersey, the establishment of Jesuit's bark (Iva frutescens), groundsel-tree (Baccharis halimifolia), black rush (Juncus roemerianus), and saltmeadow cordgrass in herbicide-treated areas appeared to limit the spread of common reed populations [236]. In ponds at Cape Cod National Seashore, repeated stem breakage in a high-water year produced substantial common reed mortality. The number of live stems decreased by 58% to 99% in treated ponds [209].

Several studies report the effects of combining herbicides with fire to reduce common reed. These studies are discussed in Fire as a control method.

Fire: See Fire Management Considerations.

Biological: While there have been no purposeful introductions of insects that target the nonnative common reed haplotype, many have been accidentally introduced. Likely they arrived in shipments packed with dried common reed material. The diversity and abundance of these herbivores is highest near New York City [25]. There has been some discussion about the introduction and use of a haplotype-specific biocontrol [27,90]. For more on insects already in the United States and potential European introductions, see [28].

Phragmites australis: REFERENCES

1. Able, Kenneth W.; Hagan, Stacy M. 2000. Effects of common reed (Phragmites australis) invasion on marsh surface macrofauna: response of fishes and decapod crustaceans. Estuaries. 23(5): 633-646. [68790]
2. Ailes, Marilyn Carol. 1993. Phragmites australis (Cav.) Trin. ex. Steud. community response to fire. Princess Anne, MD: University of Maryland, Eastern Shore. 144 p. Thesis. [68806]
3. Ailstock, M. Stephen; Norman, C. Michael; Bushmann, Paul J. 2001. Common reed Phragmites australis: control and effects upon biodiversity in freshwater nontidal wetlands. Restoration Ecology. 9(1): 49-59. [68716]
4. Allan, Philip F. 1950. Ecological bases for land use planning in Gulf Coast marshlands. Journal of Soil and Water Conservation. 5: 57-62, 85. [14612]
5. Amsberry, Lindsay; Baker, Michael A.; Ewanchuk, Patrick J.; Bertness, Mark D. 2000. Clonal integration and the expansion of Phragmites australis. Ecological Applications. 10(4): 1110-1118. [68717]
6. Anderson, Bertin W.; Ohmart, Robert D.; Meents, Julie K.; Hunter, William C. 1984. Avian use of marshes on the lower Colorado River. In: Warner, Richard E.; Hendrix, Kathleen M., eds. California riparian systems: Ecology, conservation, and productive management: Proceedings; 1981 September 17-19; Davis, CA. Berkeley, CA: University of California Press: 598-604. [5861]
7. Anderson, Kat. 1991. Wild plant management: cross-cultural examples of the small farmers of Jaumave, Mexico, and the southern Miwok of the Yosemite region. Arid Lands Newsletter. Tucson, AZ: The University of Arizona, Office of Arid Lands Studies. 31: 18-23. [17350]
8. Anderson, Marion Kathleen. 1993. The experimental approach to assessment of the potential ecological effects of horticultural practices by indigenous peoples on California wildlands. Berkeley, CA: University of California. 211 p. Dissertation. [33081]
9. Andrews, Ross L.; Broome, Stephen W. 2006. Oak flat restoration on phosphate-mine spoils. Restoration Ecology. 14(2): 210-219. [63274]
10. Auclair, Allan N.; Bouchard, Andre; Pajaczkowski, Josephine. 1973. Plant composition and species relations on the Huntingdon Marsh, Quebec. Canadian Journal of Botany. 51: 1231-1247. [14498]
11. Baker, William L. 1984. A preliminary classification of the natural vegetation of Colorado. The Great Basin Naturalist. 44(4): 647-676. [380]
12. Banner, Roger E. 1992. Vegetation types of Utah. Journal of Range Management. 14(2): 109-114. [20298]
13. Baptista, Tony L.; Shumway, Scott W. 1998. A comparison of the seed banks of sand dunes with different disturbance histories on Cape Cod National Seashore. Rhodora. 100(903): 298-313. [65375]
14. Barkworth, Mary E.; Capels, Kathleen M.; Long, Sandy; Anderton, Laurel K.; Piep, Michael B., eds. 2007. Flora of North America north of Mexico. Volume 24: Magnoliophyta: Commelinidae (in part): Poaceae, part 1. New York: Oxford University Press. 911 p. Available online: [68092]
15. Barkworth, Mary E.; Capels, Kathleen M.; Long, Sandy; Piep, Michael B., eds. 2003. Flora of North America north of Mexico. Volume 25: Magnoliophyta: Commelinidae (in part): Poaceae, part 2. New York: Oxford University Press. 783 p. Available online: [68091]
16. Barrett, Nels E.; Niering, William A. 1993. Tidal marsh restoration: trends in vegetation change using a geographical information system (GIS). Restoration Ecology. 1(1): 18-28. [20797]
17. Barry, Matthew J.; Bowers, Richard; de Szalay, Ferenc A. 2004. Effects of hydrology, herbivory and sediment disturbance on plant recruitment in a Lake Erie coastal wetland. The American Midland Naturalist. 151(2): 217-232. [48374]
18. Bart, David. 1997. The use of local knowledge in understanding ecological change: a study of salt hay farmers' knowledge of Phragmites australis invasion. New Brunswick, NJ: Rutgers University. 139 p. Thesis. [69493]
19. Bart, David; Hartman, Jean Marie. 2000. Environmental determinants of Phragmites australis expansion in a New Jersey salt marsh: an experimental approach. Oikos. 89(1): 59-69. [68776]
20. Bart, David; Hartman, Jean Marie. 2002. Environmental constraints on early establishment of Phragmites australis in salt marshes. Wetlands. 22(2): 201-213. [68718]
21. Bart, David; Hartman, Jean Marie. 2003. The role of large rhizome dispersal and low salinity windows in the establishment of common reed, Phragmites australis, in salt marshes: new links to human activities. Estuaries. 26(2B): 436-443. [68720]
22. Bean, Lowell John; Saubel, Katherine Siva. 1972. Telmalpakh: Cahuilla Indian knowledge and usage of plants. Banning, CA: Malki Museum. 225 p. [35898]
23. Beatley, Janice C. 1976. Vascular plants of the Nevada Test Site and central-southern Nevada: ecologic and geographic distributions. [Washington, DC]: U.S. Energy Research and Development Administration, Office of Technical Information, Technical Information Center. 308 p. Available from U.S. Department of Commerce, National Technical Information Service, Springfield, VA. TID-26881/DAS. [63152]
24. Benoit, Lori K.; Askins, Robert A. 1999. Impact of the spread of Phragmites on the distribution of birds in Connecticut tidal marshes. Wetlands. 19(1): 194-208. [69513]
25. Blossey, B.; Schwarzlander, M.; Haflinger, P.; Casagrande, R.; Tewksbury, L. 2002. Common reed. In: Van Driesche, Roy; Lyon, Suzanne; Blossey, Bernd; Hoddle, Mark; Reardon, Richard, tech. coord. Biological control of invasive plants in the eastern United States. USDA Forest Service Publication FHTET-2002-04. [Washington, DC]: U.S. Department of Agriculture, Forest Service: 146-155. Available online: [2005, August 12]. [54244]
26. Blossey, Bernd. 2002. Morphological differences between native and introduced genotypes, [Online]. In: Phragmites: common reed: Morphological differences. Ithaca, NY: Cornell University, Cornell Cooperative Extension, Ecology and Management of Invasive Plants Program (Producer). Available: [2008, March 12]. [69727]
27. Blossey, Bernd. 2003. A framework for evaluating potential ecological effects of implementing biological control of Phragmites australis. Estuaries. 26(2B): 607-617. [68721]
28. Blossey, Bernd. 2003. Insects already introduced to North America, [Online]. In: Phragmites: common reed: Insects. Ithaca, NY: Cornell University, Cornell Cooperative Extension, Ecology and Management of Invasive Plants Program (Producer). Available: [2008, March 13]. [69726]
29. Boelter, Don H.; Verry, Elon S. 1977. Peatland and water in the northern Lake States. Gen. Tech. Rep. NC-31. St. Paul, MN: U.S. Department of Agriculture, Forest Service, North Central Forest Experiment Station. 22 p. [8168]
30. Bolen, Eric G. 1964. Plant ecology of spring-fed salt marshes in western Utah. Ecological Monographs. 34(2): 143-166. [11214]
31. Boone, Jim L.; Furbish, C. Elaine; Turner, Kent. 1987. Control of Phragmites communis: results of burning, cutting, and covering with plastic in a North Carolina salt marsh. CPSU Technical Report 41. Athens, GA: University of Georgia, Institute of Ecology, Cooperative Park Studies Unit. 15 p. [68794]
32. Borgmann, Marian; Jonas, Jayne. 2003. The vascular plant community composition of three fens in the sandhills of Nebraska. In: Foré, Stephanie, ed. Promoting prairie: Proceedings of the 18th North American Prairie Conference; 2002 June 23-27; Kirksville, MO. Kirksville, MO: Truman State University Press: 164-173. [67090]
33. Bosakowski, Thomas. 1986. Short-eared owl winter roosting strategies. American Birds. 40(2): 237-240. [22249]
34. Bosakowski, Thomas. 1989. Observations on the evening departure and activity of wintering short-eared owls in New Jersey. Journal of Raptor Research. 23(4): 162-166. [22250]
35. Boyd, Steve. 1999. Vascular flora of the Liebre Mountains, western Transverse Ranges, California. Aliso. 18(2): 93-139. [40639]
36. Briea, Patricia. 2006. A study of Phragmites australis along an elevational gradient and seed germination response at different salinity levels. Lowell, MA: University of Massachusetts, Department of Environmental, Earth, and Atmospheric Sciences. 96 p. Thesis. [68797]
37. Burdick, David M.; Konisky, Raymond A. 2003. Determinants of expansion for Phragmites australis, common reed, in natural and impacted coastal marshes. Estuaries. 26(2B): 407-416. [68722]
38. Butler, Brett J.; Barclay, John S.; Fisher, Jeffrey P. 1999. Plant communities and flora of Robins Island (Long Island), New York. Journal of the Torrey Botanical Society. 126(1): 63-76. [62545]
39. Chabreck, Robert H. 1972. Vegetation, water and soil characteristics of the Louisiana coastal region. Bulletin 664. Baton Rouge, LA: Louisiana State University, Louisiana Agricultural Experiment Station. 72 p. [19976]
40. Chabreck, Robert H.; Palmisano, A. W. 1973. The effects of Hurricane Camille on the marshes of the Mississippi River Delta. Ecology. 54(5): 1118-1123. [68788]
41. Chamberlain, J. L. 1959. Gulf Coast marsh vegetation as food of wintering waterfowl. Journal of Wildlife Management. 23(1): 97-102. [14535]
42. Chambers, R. M.; Osgood, D. T.; Bart, D. J.; Montalto, F. 2003. Phragmites australis invasion and expansion in tidal wetlands: interactions among salinity, sulfide, and hydrology. Estuaries. 26(2B): 398-406. [68723]
43. Chambers, Randolph M.; Meyerson, Laura A.; Saltonstall, Kristin. 1999. Expansion of Phragmites australis into tidal wetlands of North America. Aquatic Botany. 64(3-4): 261-273. [68779]
44. Clark, Kennedy H. 1998. Use of prescribed fire to supplement control of an invasive plant, Phragmites australis, in marshes of southeast Virginia. In: Pruden, Teresa L.; Brennan, Leonard A., eds. Fire in ecosystem management: shifting the paradigm from suppression to prescription: Proceedings, Tall Timbers fire ecology conference; 1996 May 7-10; Boise, ID. No. 20. Tallahassee, FL: Tall Timbers Research Station: 140. [35620]
45. Clayton, W. D. 1968. The correct name of the common reed. Taxon. 17: 168-169. [16673]
46. Clewell, Andre F. 1985. Guide to the vascular plants of the Florida Panhandle. Tallahassee, FL: Florida State University Press. 605 p. [13124]
47. Comes, R. D.; Bruns, V. F.; Kelley, A. D. 1978. Longevity of certain weed and crop seeds in fresh water. Weed Science. 26(4): 336-344. [50697]
48. Cooper, James A. 1978. The history and breeding biology of the Canada geese of Marshy Point, Manitoba. Wildlife Monographs No. 61. Washington, DC: The Wildlife Society. 87 p. [18122]
49. Cowardin, Lewis M.; Carter, Virginia; Golet, Francis C.; LaRoe, Edward T. 1979. Classification of wetlands and deepwater habitats of the United States. FWS/OBS-79/31. Washington, DC: U.S. Department of the Interior, Fish and Wildlife Service. 131 p. [41938]
50. Croft, Lisa K.; Haley, Jennifer S.; Paulson, Larry J. 1990. The Lake Mead cover enhancement project: planting native vegetation creates new habitat. In: Hughes, H. Glenn; Bonnicksen, Thomas M., eds. Restoration `89: the new management challenge: Proceedings, 1st annual meeting of the Society for Ecological Restoration; 1989 January 16-20; Oakland, CA. Madison, WI: The University of Wisconsin Arboretum, Society for Ecological Restoration: 403-419. [14713]
51. Cronquist, Arthur; Holmgren, Arthur H.; Holmgren, Noel H.; Reveal, James L.; Holmgren, Patricia K. 1977. Intermountain flora: Vascular plants of the Intermountain West, U.S.A. Vol. 6: The Monocotyledons. New York: Columbia University Press. 584 p. [719]
52. Cross, Diana H.; Fleming, Karen L. 1989. Control of phragmites or common reed. Fish and Wildlife Leaflet 13.4.12. Washington, DC: U.S. Department of the Interior, Fish and Wildlife Service. 5 p. [18396]
53. Cross, Diana Harding. 1983. Wildlife habitat improvement by control of Phragmites communis with fire and herbicide. Fort Collins, CO: Colorado State University. 81 p. Thesis. [18403]
54. Curtis, John T. 1959. Aquatic communities. In: The vegetation of Wisconsin. Madison, WI: The University of Wisconsin Press: 385-401. [60531]
55. Cutshall, Jack R. 1994. SRM 806: Gulf Coast salt marsh. In: Shiflet, Thomas N., ed. Rangeland cover types of the United States. Denver, CO: Society for Range Management: 114. [67473]
56. Cutshall, Jack R. 1994. SRM 807: Gulf Coast fresh marsh. In: Shiflet, Thomas N., ed. Rangeland cover types of the United States. Denver, CO: Society for Range Management: 114-115. [67474]
57. D'Antonio, Carla; Meyerson, Laura A. 2002. Exotic plant species as problems and solutions in ecological restoration: a synthesis. Restoration Ecology. 10(4): 703-713. [43644]
58. Diggs, George M., Jr.; Lipscomb, Barney L.; O'Kennon, Robert J. 1999. Illustrated flora of north-central Texas. Sida Botanical Miscellany, No. 16. Fort Worth, TX: Botanical Research Institute of Texas. 1626 p. [35698]
59. DiTomaso, Joseph M.; Healy, Evelyn A. 2003. Aquatic and riparian weeds of the West. Publication 3421. Davis, CA: University of California, Agriculture and Natural Resources. 442 p. [48834]
60. Drawe, D. Lynn. 1994. SRM 726: Cordgrass. In: Shiflet, Thomas N., ed. Rangeland cover types of the United States. Denver, CO: Society for Range Management: 101-102. [67377]
61. Drifmeyer, J. E.; Zieman, J. C. 1979. Germination enhancement and inhibition of Distichlis spicata and Scirpus robustus seeds from Virginia. Estuaries. 2(1): 16-21. [54146]
62. Duke, James A. 1992. Handbook of edible weeds. Boca Raton, FL: CRC Press. 246 p. [52780]
63. Duncan, Wilbur H.; Duncan, Marion B. 1987. The Smithsonian guide to seaside plants of the Gulf and Atlantic coasts from Louisiana to Massachusetts, exclusive of lower peninsular Florida. Washington, DC: Smithsonian Institution Press. 409 p. [12906]
64. Elias, Thomas S.; Dykeman, Peter A. 1982. Field guide to North American edible wild plants. New York: Outdoor Life Books. 286 p. [21104]
65. Elmore, Francis H. 1944. Ethnobotany of the Navajo. Monograph Series: 1(7). Albuquerque, NM: University of New Mexico. 136 p. [35897]
66. Ewing, J. 1924. Plant successions of the brush-prairie in north-western Minnesota. Journal of Ecology. 12: 238-266. [11122]
67. Farnsworth, Elizabeth J.; Meyerson, Laura A. 2003. Comparative ecophysiology of four wetland plant species along a continuum of invasiveness. Wetlands. 23(4): 750-762. [68725]
68. Felger, R. S. 1977. Mesquite in Indian cultures of southwestern North America. In: Simpson, B. B., ed. Mesquite: Its biology in two desert ecosystems. US/IBP Synthesis Series 4. Stroudsburg, PA: Dowden, Hutchinson & Ross, Inc: 150-176. [5195]
69. Fell, Paul E.; Warren, R. Scott; Light, John K.; Rawson, Robert L., Jr.; Fairley, Sean M. 2003. Comparison of fish and macroinvertebrate use of Typha angustifolia, Phragmites australis, and treated Phragmites marshes along the lower Connecticut River. Estuaries. 26(2B): 534-551. [68726]
70. Figley, William K.; VanDruff, Larry W. 1982. The ecology of urban mallards. Wildlife Monographs No. 81. Washington, DC: The Wildlife Society. 40 p. [2041]
71. Fleming, G. P.; Coulling, P. P.; Patterson, K. D. 2005. Estuarine system, [Online]. In: The natural communities of Virginia: classification of ecological community groups. Second approximation. Version 2.1. Richmond, VA: Virginia Department of Conservation and Recreation, Division of Natural Heritage (Producer). Available: [2005, November 3]. [60511]
72. Flora of North America Association. 2008. Flora of North America: The flora, [Online]. Flora of North America Association (Producer). Available: [36990]
73. Francis, John K. 2004. Phragmites australis. In: Francis, John K., ed. Wildland shrubs of the United States and its territories: thamnic descriptions: volume 1. Gen. Tech. Rep. IITF-GTR-26. San Juan, PR: U.S. Department of Agriculture, Forest Service, International Institute of Tropical Forestry; Fort Collins, CO: U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station: 555-557. [52217]
74. Frieswyk, Christin B.; Zedler, Joy B. 2006. Do seed banks confer resilience to coastal wetlands invaded by Typha xglauca? Canadian Journal of Botany. 84(12): 1882-1893. [68093]
75. Frolik, A. L. 1941. Vegetation on the peat lands of Dane County, Wisconsin. Ecological Monographs. 11(1): 117-140. [16805]
76. Frost, Cecil C. 1995. Presettlement fire regimes in southeastern marshes, peatlands, and swamps. In: Cerulean, Susan I.; Engstrom, R. Todd, eds. Fire in wetlands: a management perspective: Proceedings, 19th Tall Timbers fire ecology conference; 1993 November 3-6; Tallahassee, FL. No. 19. Tallahassee, FL: Tall Timbers Research Station: 39-60. [26949]
77. Frost, Cecil Carlysle, III. 2000. Studies in landscape fire ecology and presettlement vegetation of the southeastern United States. Chaple Hill, NC: University of North Carolina. 620 p. Dissertation. [40279]
78. Gabrey, Steven W.; Afton, Alan D.; Wilson, Barry C. 2001. Effects of structural marsh management and winter burning on plant and bird communities during summer in the Gulf Coast Chenier Plain. Wildlife Society Bulletin. 29(1): 218-231. [54077]
79. Galinato, Marita Ignacio; van der Valk, A. G. 1986. Seed germination traits of annuals and emergents recruited during drawdowns in the Delta Marsh, Manitoba, Canada. Aquatic Botany. 26: 89-102. [23543]
80. Gallagher, John L.; Plumley, F. Gerald. 1979. Underground biomass profiles and productivity in Atlantic coastal marshes. American Journal of Botany. 66(2): 156-161. [68793]
81. Gates, Frank C. 1942. The bogs of northern Lower Michigan. Ecological Monographs. 12(3): 213-254. [10728]
82. Gleason, Henry A.; Cronquist, Arthur. 1991. Manual of vascular plants of northeastern United States and adjacent Canada. 2nd ed. New York: New York Botanical Garden. 910 p. [20329]
83. Good, Ralph E. 1984. Role of fire in Mid Atlantic coastal plain ecosystems. In: Foley, Mary K.; Bratton, Susan P., eds. Barrier islands: critical fire management problems: Proceedings of a workshop; 1984 February 8 - 10; Titusville, FL. [Tallahassee, FL]: [Tall Timbers Research Station]: 23. [68714]
84. Goodrich, Sherel; Neese, Elizabeth. 1986. Uinta Basin flora. Ogden, UT: U.S. Department of Agriculture, Forest Service, Intermountain Region, Ashley National Forest; Vernal, UT: U.S. Department of the Interior, Bureau of Land Management, Vernal District. 320 p. [23307]
85. Gould, Frank W. 1978. Common Texas grasses. College Station, TX: Texas A&M University Press. 267 p. [5035]
86. Gratton, Claudio; Denno, Robert F. 2006. Arthropod food web restoration following removal of an invasive wetland plant. Ecological Applications. 16(2): 622-631. [62282]
87. Great Plains Flora Association. 1986. Flora of the Great Plains. Lawrence, KS: University Press of Kansas. 1392 p. [1603]
88. Greenall, Jason Andrew. 1995. First-year regrowth of three marsh plant communities after fall and spring fires in the Delta Marsh, Manitoba. Winnipeg, MB: University of Manitoba. 122 p. Thesis. [68804]
89. Gullion, Gordon W. 1960. The ecology of Gambel's quail in Nevada and the arid Southwest. Ecology. 41(3): 518-536. [49039]
90. Hafliger, Patrick; Schwarzlander, Mark; Blossey, Bernd. 2006. Comparison of biology and host plant use of Archanara geminipuncta, Archanara dissoluta, Archanara neurica, and Arenostola phragmitidis (Lepidoptera: Noctuidae), potential biological control agents of Phramites australis (Arundineae: Poaceae). Annals of the Entomological Society of America. 99(4): 683-696. [68728]
91. Hall, David W. 1978. The grasses of Florida. Gainesville, FL: University of Florida. 498 p. Dissertation. [53560]
92. Hall, James B.; Hansen, Paul L. 1997. A preliminary riparian habitat type classification system for the Bureau of Land Management districts in southern and eastern Idaho. Tech. Bull. No. 97-11. Boise, ID: U.S. Department of the Interior, Bureau of Land Management; Missoula, MT: University of Montana, School of Forestry, Riparian and Wetland Research Program. 381 p. [28173]
93. Hann, Wendel; Havlina, Doug; Shlisky, Ayn; [and others]. 2005. Interagency fire regime condition class guidebook. Version 1.2, [Online]. In: Interagency fire regime condition class website. U.S. Department of Agriculture, Forest Service; U.S. Department of the Interior; The Nature Conservancy; Systems for Environmental Management (Producer). Variously paginated [+ appendices]. Available: [2007, May 23]. [66734]
94. Hansen, Paul L.; Chadde, Steve W.; Pfister, Robert D. 1988. Riparian dominance types of Montana. Misc. Publ. No. 49. Missoula, MT: University of Montana, School of Forestry, Montana Forest and Conservation Experiment Station. 411 p. [5660]
95. Hansen, Paul; Boggs, Keith; Pfister, Robert; Joy, John. 1990. Classification and management of riparian and wetland sites in central and eastern Montana. Draft Version 2. Missoula, MT: University of Montana, School of Forestry, Montana Forest and Conservation Experiment Station, Montana Riparian Association. 279 p. [12477]
96. Hardy, Joyce Phillips; Anderson, Val Jo; Gardner, John S. 1995. Stomatal characteristics, conductance ratios, and drought-induced leaf modifications of semiarid grassland species. American Journal of Botany. 82(1): 1-7. [25694]
97. Harrington, H. D. 1964. Manual of the plants of Colorado. 2nd ed. Chicago, IL: The Swallow Press, Inc. 666 p. [6851]
98. Harris, Stanley W.; Marshall, William H. 1960. Experimental germination of seed and establishment of seedlings of Phragmites communis. Journal of Ecology. 41: 395. [16679]
99. Haslam, S. M. 1968. The biology of reed (Phragmites communis) in relation to its control. In: Proceedings of the 9th British weed control conference; Brighton, UK: British Crop Protection Council: 392-397. [16860]
100. Haslam, S. M. 1971. Community regulation in Phragmites communis Trin. I. Monodominant stands. Journal of Ecology. 59: 65-73. [16677]
101. Haslam, S. M. 1972. Biological flora of the British Isles: Phragmites communis Trin. Journal of Ecology. 60: 585-610. [16676]
102. Haslam, S. M. 1973. Some aspects of the life history and autecology of Phragmites communis Trin. A review. Polish Archives of Hydrobiology. 20(1): 79-100. [17261]
103. Haslam, Sylvia M. 1969. The development and emergence of buds in Phragmites communis Trin. Annals of Botany. 33: 289-301. [16685]
104. Haslam, Sylvia M. 1971. The development and establishment of young plants of Phragmites communis Trin. Annals of Botany. 35: 1059-1072. [16680]
105. Havens, Kirk J.; Berquist, Harry; Priest, Walter I., III. 2003. Common reed grass, Phramites australis, expansion into constructed wetlands: are we mortgaging our wetland future? Estuaries. 26(2B): 417-422. [68729]
106. Havens, Kirk J.; Priest, Walter I., III; Berquist, Harry. 1997. Investigation and long-term monitoring of Phragmites australis within Virginia's constructed wetland sites. Environmental Management. 21(4): 599-605. [69492]
107. Hayden, Ada. 1919. The ecologic subterranean anatomy of some plants of a prairie province in central Iowa. American Journal of Botany. 6(3): 87-105. [66943]
108. Heinselman, M. L. 1970. Landscape evolution, peatland types and the environment in the Lake Agassiz Peatlands Natural Area, Minnesota. Ecological Monographs. 40(2): 235-261. [8378]
109. Hellings, Samuel E.; Gallagher, John L. 1992. The effects of salinity and flooding on Phragmites australis. Journal of Applied Ecology. 29(1): 41-49. [68781]
110. Herrick, Bradley M.; Wolf, Amy T. 2005. Invasive plant species in diked vs. undiked Great Lakes wetlands. Journal of Great Lakes Research. 31(3): 277-278. [68541]
111. Hickman, James C., ed. 1993. The Jepson manual: Higher plants of California. Berkeley, CA: University of California Press. 1400 p. [21992]
112. Hitchcock, C. Leo; Cronquist, Arthur. 1973. Flora of the Pacific Northwest. Seattle, WA: University of Washington Press. 730 p. [1168]
113. Hoagland, Bruce. 2000. The vegetation of Oklahoma: a classification for landscape mapping and conservation planning. The Southwestern Naturalist. 45(4): 385-420. [41226]
114. Holm, LeRoy G.; Plocknett, Donald L.; Pancho, Juan V.; Herberger, James P. 1977. The world's worst weeds: distribution and biology. Honolulu, HI: University Press of Hawaii. 609 p. [20702]
115. Holroyd, Edmond W., III; Eberts, Debra. 2000. Aerial documentation of effective biocontrol of purple loosestrife at Winchester Wasteway, Washington. In: Interntional conference on riparian ecology and management in multi-land use watersheds: Proceedings, American Water Resources Association's 2000 summer specialty conference; 2000 August 28-31; Portland, OR. Technical Publication Series No. TPS 00-2. Middleburg, VA: American Water Resources Association: 35-40. [37557]
116. Howard, Rebecca J.; Rafferty, Patricia S. 2006. Clonal variation in response to salinity and flooding stress in four marsh macrophytes of the northern Gulf of Mexico, USA. Environmental and Experimental Botany. 56(3): 301-313. [68543]
117. Hudon, Christiane; Gagnon, Pierre; Jean, Martin. 2005. Hydrological factors controlling the spread of common reed (Phragmites australis) in the St. Lawrence River (Quebec, Canada). EcoScience. 12(3): 347-357. [68767]
118. Hunter, Karen L.; Fox, Dewayne A.; Brown, Lori M.; Able, Kenneth W. 2006. Responses of resident marsh fishes to stages of Phragmites australis invasion in three mid Atlantic estuaries. Estuaries and Coasts. 29(3): 487-498. [68730]
119. Ishii, Jun; Kadono, Yasuro. 2002. Factors influencing seed production of Phragmites australis. Aquatic Botany. 72(2): 129-141. [41574]
120. Jacobson, George L., Jr.; Almquist-Jacobson, Heather; Winne, J. Chris. 1991. Conservation of rare plant habitat: insights from the recent history of vegetation and fire at Crystal Fen, northern Maine, USA. Biological Conservation. 57(3): 287-314. [16533]
121. Jankovsky-Jones, Mabel; Rust, Steven K.; Moseley, Robert K. 1999. Riparian reference areas in Idaho: a catalog of plant associations and conservation sites. Gen. Tech. Rep. RMRS-GTR-20. Ogden, UT: U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station. 141 p. [29900]
122. Jenkins, Robert. 1973. Ecosystem restoration. In: Hulbert, Lloyd C., ed. Third Midwest prairie conference proceedings; 1972 September 22-23; Manhattan, KS. Manhattan, KS: Kansas State University, Division of Biology: 23-27. [18794]
123. Johnson, A. Sydney; Hillestad, Hilburn O.; Shanholtzer, Sheryl Fanning; Shanholtzer, G. Frederick. 1974. An ecological survey of the coastal region of Georgia. Scientific Monograph Series No. 3. Washington, DC: U.S. Department of the Interior, National Park Service. 233 p. [16102]
124. Johnston, Barry C. 1987. Plant associations of Region Two: Potential plant communities of Wyoming, South Dakota, Nebraska, Colorado, and Kansas. 4th ed. R2-ECOL-87-2. Lakewood, CO: U.S. Department of Agriculture, Forest Service, Rocky Mountain Region. 429 p. [54304]
125. Jonsson-Ninniss, Susan; Middleton, John. 1991. Effect of peat extraction on the vegetation in Wainfleet Bog, Ontario. The Canadian Field-Naturalist. 105(4): 505-511. [19716]
126. Kartesz, John T. 1999. A synonymized checklist and atlas with biological attributes for the vascular flora of the United States, Canada, and Greenland. 1st ed. In: Kartesz, John T.; Meacham, Christopher A. Synthesis of the North American flora (Windows Version 1.0), [CD-ROM]. Chapel Hill, NC: North Carolina Botanical Garden (Producer). In cooperation with: The Nature Conservancy; U.S. Department of Agriculture, Natural Resources Conservation Service; U.S. Department of the Interior, Fish and Wildlife Service. [36715]
127. Kartesz, John Thomas. 1988. A flora of Nevada. Reno, NV: University of Nevada. 1729 p. [In 2 volumes]. Dissertation. [42426]
128. Kearney, Thomas H.; Peebles, Robert H.; Howell, John Thomas; McClintock, Elizabeth. 1960. Arizona flora. 2nd ed. Berkeley, CA: University of California Press. 1085 p. [6563]
129. Keller, Barbara E. M. 2000. Genetic variation among and within populations of Phragmites australis in the Charles River watershed. Aquatic Botany. 66(3): 195-208. [68777]
130. King, Ryan S.; Deluca, William V.; Whigham, Dennis F.; Marra, Peter P. 2007. Threshold effects of coastal urbanization on Phragmites australis (common reed) abundance and foliar nitrogen in Chesapeake Bay. Estuaries and Coasts. 30(3): 469-481. [68735]
131. Kinler, Noel W.; Linscombe, Greg; Ramsey, Paul R. 1987. Nutria. In: Novak, Milan; Baker, James A.; Obbard, Martyn E.; Malloch, Bruce, eds. Wild furbearer management and conservation in North America. North Bay, ON: Ontario Trappers Association: 326-342. [50675]
132. Kirby, Donald R.; Green, Douglas M.; Mings, Thomas S. 1989. Nutrient composition of selected emergent macrophytes in northern prairie wetlands. Journal of Range Management. 42: 323-326. [6802]
133. Kiviat, Erik. 1987. Common reed (Phragmites australis). In: Decker, Daniel J.; Enck, Jody W., eds. Exotic plants with identified detrimental impacts on wildlife habitats in New York state. Natural Resources Research and Extension Series 29. Ithaca, NY: New York Chapter, Wildlife Society. 22-30. [20396]
134. Konisky, Raymond A.; Burdick, David M. 2004. Effects of stressors on invasive and halophytic plants of New England salt marshes: a framework for predicting response to tidal restoration. Wetlands. 24(2): 434-447. [68736]
135. Kruse, Arnold D.; Higgins, Kenneth F. 1998. Effects of prescribed fire upon wildlife habitat in northern mixed-grass prairie. In: Alexander, M. E.; Bisgrove, G. F., tech. coords. The art and science of fire management: Proceedings of the 1st Interior West Fire Council annual meeting and workshop; 1988 October 24-27; Kananaskis Village, AB. Information Report NOR-X-309. Edmonton, AB: Forestry Canada, Northwest Region, Northern Forestry Centre: 182-193. [40285]
136. Lackschewitz, Klaus. 1991. Vascular plants of west-central Montana--identification guidebook. Gen. Tech. Rep. INT-227. Ogden, UT: U.S. Department of Agriculture, Forest Service, Intermountain Research Station. 648 p. [13798]
137. Lambert, Adam M.; Casagrande, Richard A. 2006. Distribution of native and exotic Phragmites australis in Rhode Island. Northeastern Naturalist. 13(4): 551-560. [68763]
138. Lambert, Adam M.; Casagrande, Richard A. 2007. Characteristics of a successful estuarine invader: evidence of self-compatibility in native and non-native lineages of Phragmites australis. Marine Ecology Progress Series. 337: 299-301. [68759]
139. Lambert, Adam Matthew. 2005. Native and exotic Phragmites australis in Rhode Island: distribution and differential resistance to insect herbivores. Kingston, RI: University of Rhode Island. 106 p. Dissertation. [68799]
140. LANDFIRE Rapid Assessment. 2005. Reference condition modeling manual (Version 2.1), [Online]. In: LANDFIRE. Cooperative Agreement 04-CA-11132543-189. Boulder, CO: The Nature Conservancy; U.S. Department of Agriculture, Forest Service; U.S. Department of the Interior (Producers). 72 p. Available: /downloadfile.php?file=RA_Modeling_Manual_v2_1.pdf [2007, May 24]. [66741]
141. LANDFIRE Rapid Assessment. 2007. Rapid assessment reference condition models, [Online]. In: LANDFIRE. U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station, Fire Sciences Lab; U.S. Geological Survey; The Nature Conservancy (Producers). Available: /models_EW.php [2008, April 18] [66533]
142. Larson, Gary E. 1993. Aquatic and wetland vascular plants of the Northern Great Plains. Gen. Tech. Rep. RM-238. Fort Collins, CO: U.S. Department of Agriculture, Forest Service, Rocky Mountain Forest and Range Experiment Station. 681 p. Jamestown, ND: Northern Prairie Wildlife Research Center (Producer). Available: [2006, February 11]. [22534]
143. Lathrop, Richard G.; Windham, Lisamarie; Montesano, Paul. 2003. Does Phragmites expansion alter the structure and function of marsh landscapes? Patterns and processes revisited. Estuaries. 26(2B): 423-435. [68737]
144. League, Michael T.; Colbert, Erin P.; Seliskar, Denise M.; Gallagher, John L. 2006. Rhizome growth dynamics of native and exotic halotypes of Phragmites australis (common reed). Estuaries and Coasts. 29(2): 269-276. [68738]
145. Leithead, Horace L.; Yarlett, Lewis L.; Shiflet, Thomas N. 1971. 100 native forage grasses in 11 southern states. Agric. Handb. 389. Washington, DC: U.S. Department of Agriculture, Forest Service. 216 p. [17551]
146. Lelong, Benjamin; Lavoie, Claude; Jodoin, Yvon; Belzile, Francois. 2007. Expansion pathways of the exotic common reed (Phragmites australis): a historical and genetic analysis. Diversity and Distributions. 13(4): 430-437. [68758]
147. Leonard, Lynn A.; Wren, P. Ansley; Beavers, Rebecca L. 2002. Flow dynamics and sedimentation in Spartina alterniflora and Phragmites australis marshes of the Chesapeake Bay. Wetlands. 22(2): 415-424. [68739]
148. Looman, J. 1981. The vegetation of the Canadian prairie provinces. III. Aquatic and semi-aquatic vegetation. Phytocoenologia. 9(4): 473-497. [18401]
149. Looman, J. 1982. The vegetation of the Canadian prairie provinces. III. Aquatic and semi-aquatic vegetation. Part 2: Freshwater marshes and bogs. Phytocoenologia. 10(4): 401-423. [18402]
150. Love, Askell; Love, Doris. 1954. Vegetation of a prairie marsh. Bulletin of the Torrey Botanical Club. 81(1): 16-34. [18103]
151. Ludwig, David F.; Iannuzzi, Timothy J.; Esposito, Anthony N. 2003. Phragmites and environmental management: a question of values. Estuaries. 26(2B): 624-630. [68740]
152. Lynch, John J.; O'Neil, Ted; Lay, Daniel W. 1947. Management significance of damage by geese and muskrats to Gulf Coast marshes. Journal of Wildlife Management. 11(1): 50-76. [14559]
153. Magee, Dennis W. 1981. Freshwater wetlands: A guide to common indicator plants of the Northeast. Amherst, MA: University of Massachusetts Press. 245 p. [14824]
154. Maheu-Giroux, Mathieu; de Blois, Sylvie. 2007. Landscape ecology of Phragmites australis invasion in networks of linear wetlands. Landscape Ecology. 22(2): 285-301. [68764]
155. Marks, Marianne; Lapin, Beth; Randall, John. 1994. Phragmites australis (P. communis): threats, management, and monitoring. Natural Areas Journal. 14(4): 285-294. [26678]
156. Martin, Alex C.; Erickson, Ray C.; Steenis, John H. 1957. Improving duck marshes by weed control. Circular 19 (Revised). Washington, DC: U.S. Department of the Interior, Bureau of Sport Fisheries and Wildlife. 60 p. [16324]
157. Martin, Alexander C.; Hotchkiss, Neil; Uhler, Francis M.; Bourn, Warren S. 1953. Classification of wetlands of the United States. Special Scientific Report Wildlife No. 20. Washington, DC: U.S. Department of the Interior, Fish and Wildlife Service. 14 p. [41944]
158. Martin, William C.; Hutchins, Charles R. 1981. A flora of New Mexico. Volume 2. Germany: J. Cramer. 2589 p. [37176]
159. Mason, Herbert L. 1957. A flora of the marshes of California. Berkeley, CA: University of California Press. 878 p. [16905]
160. McGlynn, Catherine Ann. 2006. The effects of two invasive plants on native communities in Hudson River freshwater tidal wetlands. Stony Brook, NY: Stony Brook University. 218 p. Dissertation. [68567]
161. Meadows, Robert E.; Saltonstall, Kristin. 2007. Distribution of native and introduced Phragmites australis in freshwater and oligohaline tidal marshes of the Delmarva Peninsula and southern New Jersey. Journal of the Torrey Botanical Society. 134(1): 99-107. [68742]
162. Merendino, M. Todd; Smith, Loren M.; Murkin, Henry R.; Pederson, Roger L. 1990. The response of prairie wetland vegetation to seasonality of drawdown. Wildlife Society Bulletin. 18(3): 245-251. [17645]
163. Minchinton, Todd E. 2002. Disturbance by wrack facilitates spread of Phragmites australis in a coastal marsh. Journal of Experimental Marine Biology and Ecology. 281(1-2): 89-107. [68772]
164. Minchinton, Todd E. 2002. Precipitation during El Nino correlates with increasing spread of Phragmites australis in New England, USA, coastal marshes. Marine Ecology Progress Series. 242: 305-309. [68773]
165. Minchinton, Todd E.; Bertness, Mark D. 2003. Disturbance-mediated competition and the spread of Phragmites australis in a coastal marsh. Ecological Applications. 13(5): 1400-1416. [68743]
166. Minchinton, Todd E.; Simpson, Juliet C.; Bertness, Mark D. 2006. Mechanisms of exlusion of native coastal marsh plants by an invasive grass. Journal of Ecology. 94(2): 342-354. [61512]
167. Minckley, W. L.; Brown, David E. 1982. Wetlands. In: Brown, David E., ed. Biotic communities of the American Southwest--United States and Mexico. Desert Plants. 4(1-4): 223-287. [8898]
168. Mohlenbrock, Robert H. 1986. [Revised edition]. Guide to the vascular flora of Illinois. Carbondale, IL: Southern Illinois University Press. 507 p. [17383]
169. Munz, Philip A.; Keck, David D. 1973. A California flora and supplement. Berkeley, CA: University of California Press. 1905 p. [6155]
170. Nevada Natural Heritage Program. 2003. National vegetation classification for Nevada [NVC], [Online]. Carson City, NV: Nevada Department of Conservation and Natural Resources (Producer). Available: [2005, November 3]. [55021]
171. O'Neil, Ted. 1949. The muskrat in the Louisiana coastal marshes. New Orleans, LA: Louisiana Department of Wildlife and Fisheries, Fish and Game Division, Federal Aid Section. 152 p. [18182]
172. Otto, Sibylle; Groffman, Peter M.; Findlay, Stuart E. G.; Arreola, Anna E. 1999. Invasive plant species and microbial processes in a tidal freshwater marsh. Journal of Environmental Quality. 28(4): 1252-1257. [37547]
173. Pace, Wm. Lynn, III; Riskind, David H.; Hayes, Tom D. 1988. Restoration and management of native plant communities on Texas parklands: the mixed-prairie experience. In: Davis, Arnold; Stanford, Geoffrey, eds. The prairie: roots of our culture; foundation of our economy: Proceedings of the 10th North American prairie conference; 1986 June 22-26; Denton, TX. Dallas, TX: Native Prairie Association of Texas: 09.04: 1-3. [25605]
174. Parsons, Katherine C. 2003. Reproductive success of wading birds using Phragmites marsh and upland nesting habitats. Estuaries. 26(2B): 596-601. [68744]
175. Payne, Richard E.; Blossey, Bernd. 2007. Presence and abundance of native and introduced Phragmites australis (Poaceae) in Falmouth, Massachusetts. Rhodora. 109(937): 96-100. [68760]
176. Pearcy, R. W.; Berry J. A.; Bartholomew, B. 1974. Field photosynthetic performance and leaf temperature of Phragmites communis under summer conditions in Death Valley, California. Photosynthetica. 8(2): 104-108. [69699]
177. Pederson, Roger L. 1981. Seed bank characteristics of the Delta Marsh, Manitoba: applications for wetland management. In: Richardson, B., ed. Midwest conference on wetland values and management: Selected proceedings; 1981 June 17-19; St. Paul, MN. Minneapolis, MN: Freshwater Society: 61-69. [24016]
178. Penfound, W. T.; Hathaway, Edward S. 1938. Plant communities in the marshlands of southeastern Louisiana. Ecological Monographs. 8(1): 3-56. [15089]
179. Philipp, Charles C.; Brown, Russell G. 1965. Ecological studies of transition-zone vascular plants in South River, Maryland. Chesapeake Science. 6(2): 73-81. [62591]
180. Philipp, Kurt R.; Field, Richard T. 2005. Phragmites australis expansion in Delaware Bay salt marshes. Ecological Engineering. 25(3): 275-291. [68766]
181. Pojar, Jim; MacKinnon, Andy, eds. 1994. Plants of the Pacific Northwest coast: Washington, Oregon, British Columbia and Alaska. Redmond, WA: Lone Pine Publishing. 526 p. [25159]
182. Radford, Albert E.; Ahles, Harry E.; Bell, C. Ritchie. 1968. Manual of the vascular flora of the Carolinas. Chapel Hill, NC: The University of North Carolina Press. 1183 p. [7606]
183. Raichel, Diana L.; Able, Kenneth W.; Hartman, Jean Marie. 2003. The influence of Phragmites (common reed) on the distribution, abundance, and potential prey of a resident marsh fish in the Hackensack Meadowlands, New Jersey. Estuaries. 26(2B): 511-521. [68745]
184. Randall, John M. 1995. Weeds and natural areas management. In: Brenton, Robert; Sherlock, Joe, tech. coords. Proceedings: 16th annual forest vegetation management conference; 1995 January 10-12; Sacramento, CA. Redding, CA: Forest Vegetation Management Conference: 23-28. [27750]
185. Raunkiaer, C. 1934. The life forms of plants and statistical plant geography. Oxford: Clarendon Press. 632 p. [2843]
186. Rice, Dan; Rooth, Jill; Stevenson, J. Court. 2000. Colonization and expansion of Phragmites australis in upper Chesapeake Bay tidal marshes. Wetlands. 20(2): 280-299. [68746]
187. Richburg, Julie A.; Patterson, William A., III; Lowenstein, Frank. 2001. Effects of road salt and Phragmites australis invasion on the vegetation of a western Massachusetts calcareous lake-basin fen. Wetlands. 21(2): 247-255. [68747]
188. Robertson, Morgan M. 1997. Prescribed burning as a management and restoration tool in wetlands of the upper Midwest. In: Restoration and reclamation review: Student on-line journal (Hort 5015/5071): Vol. 2-spring 1997: restoration techniques. Available: [2007, December 18]. [68900]
189. Rogers, Garry F.; Robertson, John M.; Solecki, William D.; Vint, Mary K. 1985. Rate of Myrica pensylvanica (bayberry) expansion in grassland at Gateway National Recreation Area, New York. Bulletin of the Torrey Botanical Club. 112(1): 74-78. [39655]
190. Roland, A. E.; Smith, E. C. 1969. The flora of Nova Scotia. Halifax, NS: Nova Scotia Museum. 746 p. [13158]
191. Rooth, J. E.; Stevenson, J. C. 2000. Sediment deposition patterns in Phragmites australis communities: implications for coastal areas threatened by rising sea-level. Wetlands Ecology and Management. 8: 173-183. [68751]
192. Rooth, Jill E.; Stevenson, J. Court; Cornwell, Jeffrey C. 2003. Increased sediment accretion rates following invasion by Phragmites australis: the role of litter. Estuaries. 26(2B): 475-483. [68748]
193. Rowman, Charles T.; Daiber, Franklin C. 1984. Aboveground and belowground primary production dynamics of two Delaware Bay tidal marshes. Bulletin of the Torrey Botanical Club. 111(1): 34-41. [68789]
194. Rydberg, P. A. 1915. Phytogeographical notes on the Rocky Mountain region V. Grasslands of the subalpine and montane zones. Bulletin of the Torrey Botanical Club. 42(11): 629-642. [60596]
195. Saltonstall, Kristin. 2002. Cryptic invasion by a non-native genotype of the common reed, Phragmites australis, into North America. Proceedings of the National Academy of Sciences of the United States of America. 99(4): 2445-2449. [68774]
196. Saltonstall, Kristin. 2003. Genetic variation among North American populations of Phragmites australis: implications for management. Estuaries. 26(2B): 444-451. [68749]
197. Saltonstall, Kristin; Peterson, Paul M.; Soreng, Robert J. 2004. Recognition of Phragmites australis subsp. americanus (Poaceae: Arundinoideae) in North America: evidence from morphological and genetic analyses. SIDA. 21(2): 683-692. [69716]
198. Saltonstall, Kristin; Stevenson, J. Court. 2007. The effects of nutrients on seedling growth of native and introduced Phragmites australis. Aquatic Botany. 86(4): 331-336. [68761]
199. Schlichtemeier, Gary. 1967. Marsh burning for waterfowl. In: Proceedings, 6th annual Tall Timbers fire ecology conference; 1967 March 6-7; Tallahassee, FL. No. 6. Tallahassee, FL: Tall Timbers Research Station: 40-46. [16450]
200. Schneider, Rick E.; Faber-Langendoen, Don; Crawford, Rex C.; Weakley, Alan S. 1997. The status of biodiversity in the Great Plains: Great Plains vegetation classification. Supplemental Document 1. In: Ostlie, Wayne R.; Schneider, Rick E.; Aldrich, Janette Marie; Faust, Thomas M.; McKim, Robert L. B.; Chaplin, Stephen J., compilers. The status of biodiversity in the Great Plains, [Online]. Arlington, VA: The Nature Conservancy (Producer). 75 p. Available: [2006, May 16]. On file with: U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station, Fire Sciences Laboratory, Missoula, MT. [62020]
201. Seymour, Frank Conkling. 1982. The flora of New England. 2nd ed. Phytologia Memoirs 5. Plainfield, NJ: Harold N. Moldenke and Alma L. Moldenke. 611 p. [7604]
202. Shay, Jennifer M.; Shay, C. Thomas. 1986. Prairie marshes in western Canada, with specific reference to the ecology of five emergent macrophytes. Canadian Journal of Botany. 64: 443-454. [18397]
203. Sherff, E. E. 1912. The vegetation of Skokie Marsh, with special reference to subterranean organs and their interrelationships. Botanical Gazette. 53(5): 415-435. [66922]
204. Silliman, Brian R.; Bertness, Mark D. 2004. Shoreline development drives invasion of Phragmites australis and the loss of plant diversity on New England salt marshes. Conservation Biology. 18(5): 1424-1434. [68753]
205. Smeins, F. E.; Diamond, D. D.; Hanselka, C. W. 1992. Coastal prairie. In: Coupland, R. T., ed. Natural grasslands: Introduction and western hemisphere. Ecosystems of the World 8A. Amsterdam, Netherlands: Elsevier Science Publishers B. V.: 269-290. [23828]
206. Smith, Loren M.; Kadlec, John A. 1983. Seed banks and their role during drawdown of a North American marsh. Journal of Applied Ecology. 20: 673-684. [11196]
207. Smith, Loren Michael. 1983. Effects of prescribed burning on the ecology of a Utah marsh. Logan, UT: Utah State University. 159 p. Dissertation. [10218]
208. Smith, Robert H. 1942. Management of salt marshes on the Atlantic Coast of the United States. Transactions, 7th North American Wildlife Conference. 7: 272-277. [14505]
209. Smith, Stephen M. 2005. Manual control of Phragmites australis in freshwater ponds of Cape Cod National Seashore, Massachusetts, USA. Journal of Aquatic Plant Management. 43: 50-53. [68768]
210. Stallard, Harvey. 1929. Secondary succession in the climax forest formations of northern Minnesota. Ecology. 10(4): 476-547. [3808]
211. Stalter, Richard; Baden, John. 1994. A twenty year comparison of vegetation of three abandoned rice fields, Georgetown County, South Carolina. Castanea. 59(1): 69-77. [26043]
212. Stanton, Lee Ellis. 2005. The establishment, expansion and ecosystem effects of Phragmites australis, an invasive species in coastal Louisiana. Baton Rouge, LA: Louisiana State University, Agricultural and Mechanical College. 166 p. Dissertation. [68800]
213. Stevens, Lawrence E.; Schmidt, John C.; Ayers, Tina J.; Brown, Bryan T. 1995. Flow regulation, geomorphology, and Colorado River marsh development in the Grand Canyon, Arizona. Ecological Applications. 5(4): 1025-1039. [48984]
214. Stickney, Peter F. 1989. Seral origin of species comprising secondary plant succession in Northern Rocky Mountain forests. FEIS workshop: Postfire regeneration. Unpublished draft on file at: U.S. Department of Agriculture, Forest Service, Intermountain Research Station, Fire Sciences Laboratory, Missoula, MT. 10 p. [20090]
215. Strausbaugh, P. D.; Core, Earl L. 1977. Flora of West Virginia. 2nd ed. Morgantown, WV: Seneca Books, Inc. 1079 p. [23213]
216. Stubbendieck, James; Coffin, Mitchell J.; Landholt, L. M. 2003. Weeds of the Great Plains. 3rd ed. Lincoln, NE: Nebraska Department of Agriculture, Bureau of Plant Industry. 605 p. In cooperation with: University of Nebraska, Lincoln. [50776]
217. Suring, Lowell H.; Rowland, Mary M.; Wisdom, Michael J.; Schueck, Linda; Meinke, Cara W. 2005. Chapter 3: vegetation communities. In: Wisdom, Michael J.; Rowland, Mary M.; Suring, Lowell H., eds. Habitat threats in the sagebrush ecosystem: methods of regional assessment and applications in the Great Basin. Lawrence, KS: Alliance Communications Group: 94-113. [67402]
218. Swanson, George A.; Duebbert, Harold F. 1989. Wetland habitats of waterfowl in the prairie pothole region. In: van der Valk, Arnold, ed. Northern prairie wetlands. Ames, IA: Iowa State University Press: 228-267. [15218]
219. Taylor, Daniel M. 1984. Winter food habits of two sympatric owl species. Murrelet. 65(2): 48-49. [22257]
220. Templer, Pamela; Findlay, Stuart; Wigand, Cathleen. 1998. Sediment chemistry associated with native and non-native emergent macrophytes of a Hudson River marsh ecosystem. Wetlands. 18(1): 70-78. [37541]
221. Thompson, D. J.; Shay, J. M. 1985. The effects of fire on Phragmites australis in the Delta Marsh, Manitoba. Canadian Journal of Botany. 63: 1864-1869. [11481]
222. Thompson, D. J.; Shay, Jennifer M. 1989. First-year response of a Phragmites marsh community to seasonal burning. Canadian Journal of Botany. 67: 1448-1455. [7312]
223. Thompson, Donald James. 1981. Effects of fire on Phragmites australis (Cav.) Trin. ex Steudel and associated species at Delta Marsh, Manitoba. Winnipeg, MB: University of Manitoba. 199 p. Thesis. [51761]
224. Thorne, Robert F. 1982. The desert and other transmontane plant communities of southern California. Aliso. 10(2): 219-257. [3768]
225. Tolstead, W. L. 1942. Vegetation of the northern part of Cherry County, Nebraska. Ecological Monographs. 12: 255-292. [4470]
226. Toorn, J. van der; Mook, J. H. 1982. The influence of environmental factors and management on stands of Phragmites australis. I. Effect of burning, frost and insect damage on shoot density and shoot size. Journal of Applied Ecology. 19: 477-499. [16320]
227. Tu, Mandy, ed. 2000. Techniques from TNC stewards for the eradication of Lythrum salicaria (purple loosestrife) and Phragmites australis (common reed/Phrag) in wetlands. In: Control comments from stewards. Weeds on the web: Wildland invasive species program, [Online]. Arlington, VA: The Nature Conservancy (Producer). Available: [40056]
228. Tu, Mandy; Hurd, Callie; Randall, John M., eds. 2001. Weed control methods handbook: tools and techniques for use in natural areas. Davis, CA: The Nature Conservancy. 194 p. [37787]
229. Turner, R. E.; Warren, R. S. 2003. Valuation of continuous and intermittent Phragmites control. Estuaries. 26(2B): 618-623. [68754]
230. U.S. Department of Agriculture, Forest Service, Eastern Region. 2004. Eastern Region invasive plants ranked by degree of invasiveness, [Online]. In: Noxious weeds and non-native invasive plants. Section 3: Invasive plants. Milwaukee, WI: Eastern Region (Producer). Available: /r9/wildlife/range/weed/Sec3B.htm [2004, February 16]. [46748]
231. U.S. Department of Agriculture, Natural Resources Conservation Service. 2008. PLANTS Database, [Online]. Available: /. [34262]
232. van den Broek, Tom; van Diggelen, Rudy; Bobbink, Roland. 2005. Variation in seed buoyancy of species in wetland ecosystems with different flooding dynamics. Journal of Vegetation Science. 16: 579-586. [60322]
233. Vasquez, Edward A.; Glenn, Edward P.; Brown, J. Jed; Guntenspergen, Glenn R.; Nelson, Stephen G. 2005. Salt tolerance underlies the cryptic invasion of North American salt marshes by an introduced haplotype of the common reed Phragmites australis (Poaceae). Marine Ecology Progress Series. 298: 1-8. [69715]
234. Viosca, Percy, Jr. 1931. Spontaneous combustion in the marshes of southern Louisiana. Ecology. 12(2): 439-443. [14582]
235. Voss, Edward G. 1972. Michigan flora. Part I: Gymnosperms and monocots. Bloomfield Hills, MI: Cranbrook Institute of Science; Ann Arbor, MI: University of Michigan Herbarium. 488 p. [11471]
236. Wang, Jiangbo. 2005. Sustained restoration of Phragmites-infested wetlands: a vegetation alternative to cyclic spray and burn. Newark, DE: University of Delaware. 132 p. Dissertation. [68798]
237. Ward, Edward. 1942. Phragmites management. Transactions, 7th North American Wildlife Conference. 7: 294-298. [14959]
238. Ward, P. 1968. Fire in relation to waterfowl habitat of the delta marshes. In: Proceedings, annual Tall Timbers fire ecology conference; 1968 March 14-15; Tallahassee, FL. No. 8. Tallahassee, FL: Tall Timbers Research Station: 255-267. [18932]
239. Ward, Peter. 1974. Fires in the marsh. Manitoba Nature. Summer: 16-27. [19484]
240. Warren, R. Scott; Fell, Paul E.; Grimsby, Jonna L.; Buck, Erika L.; Rilling, G. Chris; Fertik, Rachel A. 2001. Rates, patterns, and impacts of Phragmites australis expansion and effects of experimental Phragmites control on vegetation, macroinvertebrates, and fish within tidelands of the lower Connecticut River. Estuaries. 24(1): 90-107. [68755]
241. Warren, R. Scott; Fell, Paul E.; Rozsa, Ron; Brawley, A. Hunter; Orsted, Amanda C.; Olson, Eric T.; Swamy, Varun; Niering, William A. 2002. Salt marsh restoration in Connecticut: 20 years of science and management. Restoration Ecology. 10(3): 497-513. [68517]
242. Weber, William A.; Wittmann, Ronald C. 1996. Colorado flora: eastern slope. 2nd ed. Niwot, CO: University Press of Colorado. 524 p. [27572]
243. Weis, Judith S.; Weis, Peddrick. 2003. Is the invasion of the common reed, Phragmites australis, into tidal marshes of the eastern US an ecological disaster? Marine Pollution Bulletin. 46(7): 816-820. [68771]
244. Welch, Bradley A. 2001. Phragmites australis: response to wave exposure gradients, substrate characteristics, and its influence on plant species diversity in a Lake Erie coastal marsh. Columbus, OH: The Ohio State University. 156 p. Dissertation. [68802]
245. Welling, Charles H.; Pederson, Roger L.; Valk, Arnold G. van der. 1988. Recruitment from the seed bank and the development of zonation of emergent vegetation during a drawdown in a prairie wetland. Journal of Ecology. 76: 483-496. [9276]
246. Wells, B. W. 1928. Plant communities of the coastal plain of North Carolina and their successional relations. Ecology. 9(2): 230-242. [9307]
247. Welsh, Stanley L.; Atwood, N. Duane; Goodrich, Sherel; Higgins, Larry C., eds. 1987. A Utah flora. The Great Basin Naturalist Memoir No. 9. Provo, UT: Brigham Young University. 894 p. [2944]
248. Wigand, Peter Ernest. 1987. Diamond Pond, Harney County, Oregon: vegetation history and water table in the eastern Oregon desert. The Great Basin Naturalist. 47(3): 427-458. [5980]
249. Wiggins, Ira L. 1980. Flora of Baja California. Stanford, CA: Stanford University Press. 1025 p. [21993]
250. Wijte, Antonia H. B. M. 1995. The effects of salinity and oxygen availability on the seed germination and transfer of seed reserves to the developing seedlings of salt marsh plants. Newark, DE: University of Delaware. 142 p. Dissertation. [68805]
251. Wijte, Antonia H. B. M.; Gallagher, John L. 1996. Effect of oxygen availability and salinity on early life history stages of salt marsh plants. I. Different germination strategies of Spartina alterniflora and Phragmites australis (Poaceae). American Journal of Botany. 83(10): 1337-1342. [68780]
252. Wilcox, Kerrie L.; Petrie, Scott A.; Maynard, Laurie A.; Meyer, Shawn W. 2003. Historical distribution and abundance of Phragmites australis at Long Point, Lake Erie, Ontario. Journal of Great Lakes Research. 29(4): 664-680. [68757]
253. Williams, Cecil S.; Marshall, Wm. H. 1938. Duck nesting studies, Bear River Migratory Bird Refuge, Utah, 1937. Journal of Wildlife Management. 2(2): 29-52. [11191]
254. Willner, Gale R.; Chapman, Joseph A.; Pursley, Duane. 1979. Reproduction, physiological responses, food habits, and abundance of nutria on Maryland marshes. Wildlife Monographs No. 65. Washington, DC: The Wildlife Society. 43 p. [18121]
255. Windham, Lisamarie; Lathrop, Richard G., Jr. 1999. Effects of Phragmites australis (common reed) invasion on aboveground biomass and soil properties in brackish tidal marsh of the Mullica River, New Jersey. Estuaries. 22(4): 927-935. [54144]
256. Wunderlin, Richard P. 1998. Guide to the vascular plants of Florida. Gainesville, FL: University Press of Florida. 806 p. [28655]
257. Zigmond, Maurice L. 1981. Kawaiisu ethnobotany. Salt Lake City, UT: University of Utah Press. 102 p. [35936]

FEIS Home Page