Biological Invasions

, Volume 12, Issue 5, pp 1243–1252

Impacts of the invasive plant Fallopia japonica (Houtt.) on plant communities and ecosystem processes


    • Biology DepartmentUniversity of Massachusetts
  • Peter Alpert
    • Biology DepartmentUniversity of Massachusetts
  • Jeffrey S. Dukes
    • Biology DepartmentUniversity of Massachusetts
    • Department of Forestry and Natural ResourcesPurdue University
  • Robin Harrington
    • Natural Resources ConservationUniversity of Massachusetts
Original Paper

DOI: 10.1007/s10530-009-9543-z

Cite this article as:
Aguilera, A.G., Alpert, P., Dukes, J.S. et al. Biol Invasions (2010) 12: 1243. doi:10.1007/s10530-009-9543-z


Fallopia japonica (Japanese knotweed) invades riparian areas and roadsides in New England. This large clonal species drastically alters the appearance of habitats by forming highly productive near-monocultures. To understand how these invasions affect ecosystem processes in New England, we quantified the impacts of F. japonica on species diversity, primary productivity, and nitrogen cycling at five locations in central Massachusetts, USA. In stands of F. japonica and in adjacent uninvaded areas, we recorded the cover of each plant species and measured the aboveground biomass and nitrogen (N) concentrations in plants, along with N retranslocation from F. japonica leaves and several soil characteristics. In addition, we severed rhizomes of peripheral F. japonica shoots to determine if clonal integration contributes to the species’ rapid spread and dominance. Stands of F. japonica had lower species diversity, but greater aboveground biomass and standing N than uninvaded areas. Nitrogen and carbon concentrations in biomass and N mineralization rates in soil did not differ between stands and adjacent areas. Rhizome severing temporarily reduced growth of F. japonica, suggesting that retranslocation of photoassimilates and/or nutrients between shoots via rhizomatal connections may maximize stand level growth rates and facilitate dominance by F. japonica.


Fallopia japonicaJapanese knotweedDiversityNitrogen mineralizationClonal growthInvasive species


A main reason for concern about the spread of introduced species is their potential to negatively affect ecological systems, including native species and processes such as productivity and nutrient cycling. To manage invasions and preserve systems, we need to characterize the community- and ecosystem-level effects of introduced species and elucidate the traits that enable these species to have such effects. Large, clonal plant species seem to have high potential for invasiveness and community- and ecosystem- level effects. Clonal growth traits such as resource sharing may increase the abundance of these species and thus their impact on local processes.

Fallopia japonica (Japanese knotweed) is a clonal, fast-growing, herbaceous, perennial geophyte native to Asia that reaches 2 m in height. An attractive flowering plant, F. japonica was imported to the United States as early as 1877 for ornamental purposes, and now occurs widely in riparian habitats and in disturbed habitats such as along railroad tracks and on roadsides (Forman and Kesseli 2003; Bram and McNair 2004). Considered an aggressive invader in Europe, the United States, and Canada, F. japonica spreads by clonal, rhizomatous growth and can quickly form a monoculture. Stem fragments as small as a node can produce buds and shoots and give rise to new plants (De Waal 2001). While clonal growth is widely considered the major method of population expansion for F. japonica, recent studies have shown that populations in the northeastern United States produce viable seeds in the fall with high germination rates, that seedlings can over-winter (Bram and McNair 2004; Forman and Kesseli 2003), that hybridization between F. japonica and F. sachelinensis is common (Gammon et al. 2007) and that populations are composed of multiple genets likely produced by sexual reproduction (Grimsby et al. 2007).

Fallopia japonica is one of a disproportionate number of invasive plant species that exhibit clonal growth (e.g., Ammophila arenaria, Calystegia soldanella; Acosta et al. 2006; Liu et al. 2006). Many of these species form dense patches, dramatically reducing local species diversity (Dassonville et al. 2007; Hulme and Bremner 2006). This suggests that clonal growth traits may increase the impact and/or invasiveness of introduced species. Although communities with fewer species are often less productive (Hooper et al. 2005), some invasive species have relatively extended growing periods, rapid growth rates, or large size, and can both decrease diversity and increase biomass production in plant communities (Ehrenfeld 2003).

Invasive species also often exhibit differences in nitrogen use efficiency (productivity per unit N taken up or lost), and plant and litter chemistry relative to the native community. These changes have been shown to create feedbacks that either accelerate or slow nutrient cycling (Angeloni et al. 2006; Ehrenfeld 2003; Standish et al. 2003; Vanderhoeven et al. 2006). For example, Allison and Vitousek (2004) examined litter properties, decomposition rates, and nutrient dynamics for native and invasive understory plants in the Hawaiian Islands. They found that invasive plants in Hawaii had higher specific leaf area (SLA) and faster rates of decay than native plants. Among the angiosperms in the study, invasive plants had higher nutrient concentrations than natives and, among all plant groups, litter of invasives released more nutrients (N and P) than that of natives. Experimental fertilization of plots increased litter decay rates, providing evidence for a positive feedback on nutrient cycling created by these invasive plants. In other cases, invasive plants slow decomposition rates. For example, areas on the Colorado Plateau invaded by Bromus tectorum have relatively slow rates of N mineralization (Evans et al. 2001), possibly because litter of Bromus tectorum has greater C:N and lignin:N ratios than that of native species and thereby reduces nitrogen available for microbial activity.

We hypothesized that F. japonica greatly reduces diversity of natives and slows nutrient cycling where it invades natural communities of Massachusetts, USA. We further hypothesized that physiological support of vegetative offspring by connected, established F. japonica plants of the same clone facilitates these effects. F. japonica is an ideal model species in which to test for a role of clonal growth traits in invasiveness and for effects of invasion on ecosystem properties such as primary productivity, and nutrient cycling. The species has several characteristics that suggest that it might slow N cycling: (1) rapid clonal growth with efficient N and C translocation, limiting release of N back to the environment (Adachi et al. 1996; Barney et al. 2006; Beerling et al. 1994; Price et al. 2001); (2) High lignin:N ratios in litter and thus potentially slow rates of decomposition; (3) large underground storage organs capable of sequestering large amounts of N; and (4) high relative growth rates under low N conditions, which could engender a positive feedback between growth of F. japonica and reduced N cycling (Adachi et al. 1996; Chiba and Hirose 1993). Specifically, we suggest that through positive feedbacks of increased primary productivity, efficient use of nitrogen by retention of nitrogen at senescence for remobilization in the spring, and increased nitrogen storage in belowground tissues, F. japonica may be sequestering available nitrogen, slowing decomposition, reducing rates of N mineralization, and thereby making N less available. In addition, because F. japonica shows high relative growth rates when nitrogen is limiting (Adachi et al. 1996; Chiba and Hirose 1993), a reduction in nitrogen availability could potentially increase F. japonica’s relative competitive ability in some communities, facilitating its spread. The ability of F. japonica to remobilize and redistribute N may also contribute to its success. On Mt. Fuji in Japan, severing peripheral shoots from the clone reduced their shoot growth and percent nitrogen in leaves (Adachi et al. 1996). Sequestering and sharing N between shoots, thus limiting N availability to surrounding vegetation, may be one-way F. japonica gains a competitive advantage.

We sought to understand how F. japonica alters the properties and processes of the New England plant communities it invades. Specifically, the objective of this study was to test the following hypotheses: (1) F. japonica invasion reduces species richness and diversity, (2) F. japonica invasion reduces available mineral N in the soil, (3) F. japonica invasion in North American deciduous forest increases primary productivity and plant N uptake by creating massive F. japonica monocultures, and (4) clonal integration promotes expansion of F. japonica stands.


Site selection

We selected five F. japonica stands in central Massachusetts that were each at least 100 m2 and had adjacent areas with no F. japonica. We chose two stands along a small tributary to the Fort River in Amherst, (hereafter AM1 and AM2); and three stands in Northampton: two at Rainbow Beach, a riparian area along the Connecticut River (hereafter these stands will be referred to as RB1 and RB2) and one in Northampton Meadows (NHM), a mesic forested site that contains a vernal pool. The Amherst stands receive full sun and are surrounded by native and nonnative grasses and forbs. The remaining three stands are in shaded hardwood forest understory and are surrounded by native and nonnative herbaceous plants and saplings. Dominant canopy species at the RB1 and RB2 include Acer saccharinum and Populus deltoides. Dominant canopy species at NHM include Acer saccharum, Acer saccharinum, and Carya ovata. All stands were within 16 km of each other and were at least 30 m apart.


In July of 2007 we randomly located six plots, each 1 × 1 m, inside each F. japonica stand and six additional plots in the adjacent vegetation. To establish the plots outside the F.japonica stand, we defined a boundary between the F. japonica and adjacent vegetation. We then randomly located the plots in the adjacent vegetation between 2 and 6 m from this boundary. We maintained the 2 m minimum distance to minimize shading and litter fall by F. japonica, and the 6 m maximum distance to minimize any preexisting soil differences between the two vegetation types. Cover class of each plant species in each plot was visually estimated by classes: 7 (75–100%), 6 (50–75%), 5 (25–50%), 4 (10–20%), 3 (5–10%), 2 (1–5%), 1 (<1%) as used by Braun-Blanquet (1972). From Shannon diversity indices we calculated effective species richness (eH′), where
$$ H^{\prime} = - \sum\limits_{i = 1}^{S} {p_{i} \ln p_{i} } $$
and pi is the proportion of total cover made up by species i. In the forested stands we also analyzed tree sapling data to quantify knotweed’s effect on forest regeneration.

Because we compared characteristics in and outside existing stands of F. japonica, we cannot rule out the possibility that differences existed prior to invasion, rather than as a result of invasion. For example, F. japonica might tend to invade sites with certain community or ecosystem characteristics. This is a limitation of any study that does not measure the same areas before and after invasion, which would have been very difficult to arrange in this case. However, we think it unlikely that the differences we did find were mainly pre-existent and discuss this below.

Soil analysis

At each stand we established 12 soil analysis plots, each 1 × 1 m as was done for the diversity survey. We took a soil core 5 cm in diameter and 10 cm in depth from the center of each plot on October 2, 2007. Soil cores from the same stand and microhabitat (inside or outside stand) were combined and homogenized. In the lab we dried and ground a subsample from each composite at 65°C and analyzed soil for total N and C using a dry combustion C/N analyzer (Costech Elemental Combustion System, Valencia, CA). Three additional subsamples of about 10 g each from each composite were used to estimate relative N mineralization rate. One of the subsamples was weighed, dried at 105°C, and reweighed to measure water content. Another was weighed, shaken with 100 ml of 2 M KCl (potassium chloride) in a plastic cup, shaken under ambient conditions in the lab for 24 h, filtered through filter paper, and analyzed for NH4 and NO3 with a Lachat Quickchem 8500 autoanalyzer. The remaining subsample was incubated in the dark at 25°C for 4 weeks and then extracted and analyzed as described. We calculated net ammonification rate in as the difference between the incubated and the unincubated subsample in NH4+ and net mineralization rate as the difference in NH4+ plus NO3. Results are reported in μgN g−1soil day−1.

Biomass harvest

At the end of the growing season, from October 10 to 16, 2007, we harvested all aboveground biomass (F. japonica and other species) from the soil sampling plots in three of the five stands. We were unable to harvest AM2 or RB2 due to a severe storm that toppled large trees onto these stands. We separated harvested F. japonica leaves from the stems before drying, and recorded leaf and stem biomass separately for each plot (together, these made up “total biomass”). We thoroughly homogenized each sample by hand and subsampled approximately 10% of the total biomass from each plot, and ground each subsample for C and N analysis by dry combustion as described above.

To determine the amount of nitrogen being reabsorbed from leaves during senescence, we measured changes in leaf N over time. We randomly selected three F. japonica shoots from outside of each biomass harvest plot and identified the third fully extended leaf on the main stem of each. Selected leaves were bagged in mesh and marked for collection in September, October, or at senescence, such that one leaf from each shoot was collected at each plot at each collection time. Leaves collected at the same time period were pooled by stand, dried at 60°C, ground, and analyzed for C and N by dry combustion as described above. We calculated proportion retranslocated as 1 − (N concentration of senesced leaves/N concentration of green leaves).

Rhizome severing

At the beginning of the growing season, in May 2007, we randomly located nine plots, each 0.5 × 0.5 m, around the inside edge of the invasion front of each of three stands of F. japonica at the Rainbow Beach site. We then randomly assigned 1/3 of the plots around each stand to each of three treatments: excavated and severed, excavated only, and no disturbance (a control for the effect of excavation). For the excavated and severed treatment, we removed enough soil to uncover all the parental rhizomes of the shoots of F. japonica in a plot, severed each rhizome ten nodes proximal to the shoot, and replaced the soil. For the excavated treatment, we uncovered and reburied rhizomes without severing them. When rhizomes had significant lateral growth, excavation required digging well beyond the perimeter of the plot. Every 2 weeks following the start of the treatments, we measured the basal diameter, height, and number of branches on each shoot in each plot. In weeks 5 and 7 of the experiment, we sampled the third fully extended leaf from each shoot in the plots for C and N analysis. To estimate shoot biomass from shoot basal diameter, height and branch number, we harvested 20 shoots in a fourth stand, measured their basal diameter, height, and branch number, dried them to constant mass at 65°C, and regressed mass on diameter (D), height (H), and branch (B). We used a stepwise model selection method to generate the final model: \( {\text{Mass}} = -8.41+1.271D+0.095H, \) (Adjusted R-square = 0.877). We calculated total change in biomass from first to last sampling date and log-transformed values to meet assumptions of parametric tests.

Litter chemistry

To better understand the effect of F. japonica on N cycling we conducted a separate experiment in October 2006. We collected recently senesced leaves from F. japonica and the four dominant riparian canopy species from eight sites along the Sawmill River in Montague, Massachusetts. In the lab we combined all leaves of the same species and dried all material at 60°C. We sent three replicate sub-samples to Cumberland Valley Analytical, Hagerstown, MD for nitrogen and lignin analysis.

Statistical analyses

To analyze the diversity and biomass data we first calculated means for the plots in a given microhabitat at a given site. We then used the means in two-way ANOVAs with microhabitat (inside or outside a stand of F. japonica) as a fixed effect and stand as a random effect, with a separate ANOVA for each dependent variable. To compare individual means, we used post-hoc, Tukey’s HSD tests. Total soil N and N mineralization rates were analyzed with paired t-tests. To analyze data from the severance experiment, we used a two-way ANCOVA with rhizome treatment (excavated and severed, excavated only, or no excavation) as a fixed effect, stand as a random effect, estimated initial biomass of shoots in a plot as a covariate, and change in biomass (from the initiation of treatments to the conclusion of the experiment at 14 weeks) as the dependent variable. To test effect on forest regeneration we did a randomization test using sapling data from the diversity plots. We log-transformed all biomass, N and C, soil, and rhizome growth data to meet assumptions of parametric tests. With the exception of the randomization test which was done with Microsoft Excel, all statistical analyses were performed with SPSS 17.0.



Plots outside of F. japonica stands had 1.6–10 times as many species as plots inside stands (two-way ANOVA: Microhabitat, F1,4 = 15.402, P = 0.017; Stand, F4,4 = 0.536, P = 0.720, Microhabitat × Stand, F4,50 = 10.169, P < 0.001) (Fig. 1). A total of 63 species were found outside F. japonica stands, of which 78% were native. Only 13 species, 58% of which were native, were found within stands. Across stands species richness was 2.03 ± 0.42 species m−2 (mean ± SE) in stands and 6.63 ± 1.06 outside stands. Effective species richness across stands was 1.08 ± 0.049 in stands and 5.08 ± 1.15 outside (Paired t-test: t = 3.414, df = 4, P = 0.027). In the forested site NHM, fewer established saplings grew inside the F. japonica stand (Randomization Test; P = 0.04). In RB1 and RB2 (the remaining forested stands) there were no saplings inside of F. japonica stands. However, at these stands there were too few saplings outside of F. japonica in the adjacent uninvaded vegetation to be significantly different from 0 (Randomization Test: P = 0.22, P = 0.23). Finally results show F. japonica was equally likely to exclude a non-native or a native plant species (χ2 = 0.920; P = 0.338).
Fig. 1

Mean species richness (species/m2) in F. japonica stands and in adjacent vegetation (means ± SE)

Soil nutrients

There were no significant differences between soils from F. japonica and adjacent areas for total N (Paired t-test: t = 0.053, df = 4, P = 0.960), total C (Paired t-test; t = 0.494, df = 4, P = 0.647), ammonification (Paired t-test: t = −0.87, df = 4, P = 0.935), nitrification (Paired t-test: t = 1.280, df = 4, P = 0.270),or total N mineralization (Paired t-test: t = 1.528, df = 4, P = 0.201) (Table 1).
Table 1

Percentages of soil N and C, and net mineralization values from soils under F. japonica stands and under adjacent vegetation [mean (SE)]




ΔNH4+ (μg g−1 day−1)

ΔNO3 NO2 (μg g−1 day−1)

Total min. (μg g−1 day−1)

F. japonica

0.17 (0.03)

10.90 (0.42)

0.58 (0.35)

5.09 (3.22)

5.68 (1.48)


0.17 (0.04)

10.06 (0.61)

0.36 (0.07)

4.11 (2.55)

4.48 (2.49)


Aboveground biomass in F. japonica stands was 1.8–5.2 times greater than outside F. japonica stands (two-way ANOVA: Microhabitat, F1,2 = 53.339, P = 0.018; Stand, F2,2 = 304.957, P = 0.003, Microhabitat × Stand, F2,30 = 0.070, P = 0.933) (Fig. 2). In sites NHM and AM1 there were no species other than F. japonica in the F. japonica plots. At the RB1 site non- F. japonica vegetation made up 0.05% of the total biomass across all plots. Concentration of N in whole shoots of F. japonica, calculated from the N concentrations in leaves and in stems weighted by relative mass, did not differ from N concentration in native vegetation (two-way ANOVA: Microhabitat, F1,2 = 1.076, P = 0.409; Stand, F2,2 = 3.309, P = 0.232; Microhabitat × Stand, F2,30 = 1.998, P = 0.153) (Table 2). Concentrations of C in shoots did not differ between F. japonica and native vegetation either (two-way ANOVA: Microhabitat, F1,2 = 0.310, P = 0.633; Stand, F2,2 = 1.342, P = 0.427; Microhabitat × Stand, F2,30 = 0.865, P = 0.431). However, because of the high aboveground biomass of F. japonica, standing nitrogen and carbon in aboveground vegetation were about 2–6 times higher in F. japonica stands than in total aboveground vegetation outside stands (two-way ANOVA for N: Microhabitat, F1,2 = 28.274, P = 0.034; Stand, F1,2 = 85.108, P = 0.012, Microhabitat × Stand, F2,30 = 0.216, P = 0.807; two-way ANOVA for C: Microhabitat, F1,2 = 62.613, P = 0.016, Stand F1,2 = 252.398, P = 0.004, Microhabitat × Stand, F1,2 = 0.100, P = 0.925) (Table 2).
Fig. 2

Biomass (kg/m2) in F. japonica stands and in adjacent uninvaded areas (means ± SE)

Table 2

Mean dry mass and concentrations and standing stock of N and C in F. japonica stands and adjacent vegetation [mean (SE)]



Mass (g)



Standing N (g/m2)

Standing C (g/m2)


F. japonica

1741.9 (250.6)

0.89 (0.05)

47.3 (0.1)

15.6 (2.4)

824.8 (199.8)


962.0 (86.4)

0.99 (0.10)

40.6 (1.9)

9.3 (1.0)

384.6 (28.9)


F. japonica

99.2 (23.2)

2.20 (0.10)

44.6 (0.2)

2.2 (0.5)

44.4 (10.4)


34.2 (3.3)

2.08 (0.13)

44.0 (0.5)

0.7 (0.1)

15.0 (1.4)


F. japonica

734.9 (259.3)

1.18 (0.24)

38.4 (7.7)

10.1 (3.3)

339.0 (120.1)


118.7 (23.0)

1.82 (0.10)

39.9 (0.5)

2.2 (0.5)

47.4 (9.1)

Both F. japonica leaves and adjacent vegetation outside of stands had significantly higher N concentrations than F. japonica stems (two-way ANOVA: Vegetation Type, F2,4 = 11.985, P = 0.020, Stand, F2,4 = 5.105, P = 0.079, Veg Type × Stand, F4,45 = 0.827, P = 0.515; Tukey’s HSD: Leaves-Stems, P < 0.001; Adjacent-Stems, P = 0.003). Nitrogen concentration in vegetation outside stands did not differ from N concentration in F. japonica leaves (Tukey’s HSD: Leaves-Adjacent P = 0.77). F. japonica leaves contained more standing N than F. japonica stems, despite having less biomass (two-way ANOVA: Vegetation Type, F2,4 = 14.012, P = 0.016; Stand, F2,4 = 68.111, P = 0.001; Veg Type × Stand, F4,45 = 0.566, P = 0.688; Tukey’s HSD: Leaves-Stems P = 0.001) (Table 3).
Table 3

Dry mass, concentrations, and total amounts of N and C in F. japonica stems and leaves [mean (SE)]



Mass (g)



Standing N (g/plot)

Standing C (g/plot)



576.8 (34.2)

1.86 (0.04)

47.4 (0.1)

10.9 (0.7)

273.6 (16.3)


1165.1 (69.9)

0.40 (0.01)

47.3 (0.1)

4.7 (0.3)

551.2 (33.4)



45.8 (5.0)

3.64 (0.05)

44.6 (0.1)

1.7 (0.2)

20.4 (2.2)


53.5 (4.8)

1.02 (0.04)

44.6 (0.9)

0.5 (0.1)

24.0 (2.2)



271.1 (38.2)

2.23 (0.19)

37.9 (3.1)

7.0 (0.1)

124.1 (17.6)


463.7 (67.7)

0.54 (0.05)

38.6 (3.2)

3.1 (0.5)

214.9 (31.4)


In all stands, leaf N concentrations dropped from September to senescence (Paired t-test: P = 0.035). In three of the five stands, shoots retranslocated over 60% of leaf N (Table 4).
Table 4

Concentrations of N (%) (SE) in leaves of F. japonica during the end of the growing season, and proportion retranslocated, based in decrease from September to senescence








2.36 (0.02)

1.86 (0.01)

3.58 (0.03)

2.63 (0.04)

2.66 (0.01)


1.12 (0.01)

0.60 (0.03)

3.81 (0.03)

2.26 (0.24)

2.60 (0.05)


0.78 (0.02)

0.62 (0.03)

1.84 (0.07)

1.00 (0.01)

1.86 (0.01)

% Retranslocated






Rhizome severing

Shoots with cut rhizomes produced less biomass than shoots in the uncut or control treatments (two-way ANOVA: Cutting, F2,4 = 11.162, P = 0.023; Stand, F2,4 = 5.608, P = 0.069; Cutting × Stand, F4,18 = 0.474, P = 0.755; Tukey’s HSD: Cutting–Uncut, P = 0.039; Cutting–Control, P = 0.023). Aboveground biomass production of uncut and control treatments did not differ (Tukey’s HSD: Uncut–Control, P = 0.96). Immediately after cutting, shoots from cut rhizomes grew more slowly than those in the other two treatments (Fig. 3). In all treatments, growth slowed 2 weeks after the cutting date, and stopped by mid July (week 8 census), well before fall senescence. Percent leaf N did not differ among control plots (3.57 ± 0.06, 3.32 ± 0.05; mean ± SE for weeks 5 and 7 respectively) uncut plots (3.63 ± 0.04, 3.36 ± 0.07), and cut plots (3.78 ± 0.07, 3.50 ± 0.08).
Fig. 3

Plot mean (±SE) relative growth of control ( ), uncut ( ), and cut (▲) F. japonica shoots. The y axis shows relative growth calculated as (Ln(PlotMasst) − Ln(PlotMasst+2))/Ln(PlotMasst), where t is weeks since severing rhizomes. The x axis shows time since severing treatment, starting at week 0 when treatments were imposed

Litter chemistry

Fallopia japonica leaf litter had 23–85% higher lignin/N ratios than dominant native canopy species (one-way ANOVA, F = 40.513, df = 4. P < 0.001; Tukey’s HSD P < 0.003 for all F. japonica-native comparisons) (Table 5).
Table 5

Percent litter lignin and nitrogen, and lignin: nitrogen ratios for F. japonica and four common riparian canopy species


Acer rubrum

Quercus rubra

Platanus occidentalis

Carya ovata

Fallopia japonica

Lignin (%)






Nitrogen (%)













Our results strongly suggest that F. japonica can cause large changes to the communities and ecosystems it invades. Its large size, clonal growth from, and dense monocultures place it in a class of invaders that can transform ecosystems visually, structurally, and chemically. We found support for our hypotheses that F. japonica reduces diversity and increases biomass production (Fig. 2; Table 2), and that clonal connectivity increases the ability of established stands to grow further (Fig. 3). However, we did not detect changes in N mineralization rates in F. japonica stands, despite significantly increased standing aboveground N during the growing season.

Sites with F. japonica have reduced diversity relative to uninvaded areas, and results provided preliminary evidence that F. japonica can suppress forest regeneration. Its growth form is unlike any of the native herbs that it displaces in central Massachusetts, with rhizomes that can penetrate down to 2 m below ground and extend as much as 20 m laterally (Barney et al. 2006; Weston et al. 2005). We have shown that the biomass produced by F. japonica can sometimes reach 2–6 times that of the native community. However, because leaf nitrogen concentrations do not differ between F. japonica and native vegetation it seems unlikely that increased nutrient use efficiency plays a role in F. japonica’s impressive aboveground biomass production.

Fallopia japonica’s high aboveground productivity leads to relatively large amounts of N returning to the soil via litter. Prescott (2002) argues that total nutrient return by litterfall is the best predictor of how plant species influence nutrient availability. However, we did not find differences in nitrogen mineralization rates or total N levels in the top 10 cm of soil inside stands relative to adjacent uninvaded plots. Dassonville et al. (2007) measured F. japonica’s impact on nutrient availability in Belgium and showed that F. japonica increases cation and P availability (Dassonville et al. 2007; Vanderhoeven et al. 2005). As in out study Dassonville et al. (2007), saw no significant difference in total N inside and outside of F. japonica populations. However, plant available nitrogen is not determined by the total amounts in the soil at any given time, but rather by the rate at which mineral nitrogen enters the pool by organic matter decomposition (Aber and Melillo 2001). Therefore when there is no difference in total N between sites at a single snap shot in time, there may be differences in mineralization rate and as such differences in plant available N. Our lignin:nitrogen analysis showed that F. japonica leaf litter has higher lignin/N ratios than natives. Therefore, while large amounts of N dropped with the litter, the low quality of litter (Table 5) suggests that the N in that litter might cycle slowly, offsetting the increased rate of N input via litter. The results of our soil chemistry study indicate that this is not the case. Although not significant, N mineralization was most frequently greater from soil taken from inside stands. This suggests that, as argued by Prescott (2002), the large amount of N returning in litter fall has a stronger influence on N cycling than the low quality of the litter.

During our measurement period, F.japonica reabsorbed a mean of 60% of leaf nitrogen before litter fall. This value falls within the range of N retranslocation efficiencies reported for other species (Gusewell 2005; Oikawa et al. 2008; Tylova et al. 2008). Yet, because knotweed stands are taking up more total N than adjacent vegetation and have higher standing N, the total amount of plant-retained N is higher inside stands relative to outside stands. F. japonica’s underground rhizome structures may be storing this nitrogen, which could facilitate regrowth and spread in the spring. Control strategies that employ aboveground removal of F. japonica and leave dead rhizomes in the soil to decay might therefore change soil nutrient levels. Some resorption of N may have already occurred by September when we started our sampling, so our retranslocation values are conservative. Also, comparisons between stands could be misleading because timing of senescence differed among the five populations.

Fallopia japonica’s large clonal growth form clearly influences its ability to spread, dominate, and form a monoculture. Our rhizome severing study showed connectivity to the clone to be important for biomass production (Fig. 3). Our results suggest that the shoots on the peripheral and spreading edge of the F. japonica stand receive resources for growth from the large established clone. On Mount Fuji in Japan, Adachi et al. (1996) similarly found that severing shoots from the rest of a clone reduced growth. They also found, in contrast to our results, that severed ramets had lower leaf N concentrations, possibly because N was more limiting relative to light at their sites on largely bare volcanic scree. In understory habitats, growth of peripheral shoots may be supported largely by translocation of photoassimilates. Siemens and Blossey (2007) found that F. japonica’s close relative Fallopia bohemica achieves its competitive advantage by limiting light availability to other plants, not by reducing nutrient availability or through allelopathic chemicals. At the stand level, a potential mechanism for this competitive advantage is inter-ramet access to stored nitrogen and photoassimilates. Our results suggest this mechanism may allow F. japonica to grow more rapidly and achieve greater biomass than surrounding plants and limit light availability.

Comparisons between invaded and uninvaded areas after invasion cannot fully rule out the possibility that differences between areas are the causes rather than the results of patterns of invasion. However, the very large differences in community composition within and just adjacent to stands of F. japonica strongly suggest that this species can greatly change herbaceous communities. Moreover, the large differences seen in herbaceous production and standing N seem directly attributable to the growth form of the species, which is unlike those of the native, herbaceous species it appears to be replacing in temperate North America. One of the ways it differs in growth form is its combination of large size and extensive rhizomatous growth, and the clonal integration permitted by the latter may be one important mechanism for its spread.


Funding was provided by a National Science Foundation Graduate Research Fellowship and a Northeast Alliance Fellowship to AGA. We would like to thank Jesse Schreier for field assistance. We are also grateful for the constructive and thoughtful advice of the anonymous reviewers

Copyright information

© Springer Science+Business Media B.V. 2009