The relationship between plant communities and their environment is a crucial subject in plant ecology (Soliveres et al., 2015) that changes community structure and biodiversity, resulting in characteristic vegetation (Baldeck et al., 2013; Paker et al., 2014). In this regard, the effects of invasive species invasion and environmental factors at various scales combine to produce differences in community compositions (Baldeck et al., 2013; Havel et al., 2015; Soliveres et al., 2015). The species are introduced intentionally to transport the species having high yields of crops or accidentally as a byproduct of another process (Fofonoff et al., 2003; Gaertner et al., 2016). The species invasion has rapidly increased due to rapid trade, transportation, tourism, industry growth, technological advancements, and higher urbanization rates (Ullah et al., 2022). These factors make it easier for alien species to invade different areas and endorse pressure on the natural ecosystem that provides ecosystem services to the world’s rapidly expanding human population (Pauchard et al., 2009). Therefore, invasive species have been a significant area of study for scientists worldwide (Gaertner et al., 2016; Fan et al., 2017).

An invasive species may be able to invade and spread in a new habitat despite biotic and abiotic filters (Wilson et al., 2007). Environmental parameters of any area, notably temperature, relative humidity, day length, and geographical and topographic variables, operate as an abiotic filter for newly invading invaders (Begum et al., 2021). The filter’s efficiency is determined by how closely environmental conditions correspond to the species’ native and invaded habitats (Thuiller et al., 2005). Invasive species establishment influences environmental conditions and natural plant communities directly or indirectly, including soil nutrients and physical and chemical characteristics (Weidenhamer and Callaway, 2010). Invasive plants affect evapotranspiration, causing severe threats to sustainable water supply (Levine et al., 2003) and ecosystem services such as resources and energy (Eviner et al., 2012). Invasive species may have deleterious impacts on native populations via several mechanisms, including genetic effects, pathogen introduction, and habitat degradation (Lee, 2002; Alpert, 2006).

Invasive species alter soil nutrients, particularly phosphorus (P), nitrogen (N), and potassium (K), which have a significant impact on how the soil and plants interact (Sun et al., 2019; Yinglan et al., 2019). Many studies have shown a relationship between soil N, P, and K levels and the diversity of plant species, and unimodal patterns between N, P, and K have emerged as the conventional wisdom (Pausas and Austin, 2001; John et al., 2007; Merunková and Chytrý, 2012). Species richness did not significantly alter after exogenous applications of N and P to Rengen Grassland, according to (Hejcman et al., 2010). However, it is essential to note that P and K fertilizer applications significantly changed the species composition (Hejcman et al., 2010). The total NPK contents of the soil considerably encourage the establishment of non-native shrubs and plants in the Yilong Lake Basin of the Yungui Plateau in China (Cui et al., 2009). In the hilly Ili River Valley area of China, Xu et al. (2011) revealed that the levels of TNTN, TKTK, and available phosphorus (AP) in the soil had a noticeably favorable impact on the local flora variety (Xu et al., 2011). In some instances, the soil’s pH may indicate how many nutrients are present, which can change the variety or composition of the flora (Ravolainen et al., 2013). Although there is no discernible relationship between the richness of shrubs and trees and soil pH, annual herbs, perennial herbs, and semi-shrub plants rise linearly with soil pH (Heikkinen and Neuvonen, 1997). In addition, invasive species alter the soil organic carbon (SOC), N, and pH, changing community structure and composition(Zuo et al., 2012; Cardoso et al., 2013).

However, how invasive plant communities react to floristic and vegetative characteristics and environmental factors is still being determined. Several studies (McNeish & McEwan, 2016; Subedi et al., 2019; Ullah et al., 2021; Ullah et al., 2022) strongly emphasize how invasive species traits related to environmental factors and some research focuses on how invasive species affect native plant ecosystems (Maron & Marler, 2007; Beaury et al., 2020). However, the interaction of non-native communities with one another has received very little attention in the study area, where mainly the research focuses on the invasion of single species (Khan et al., 2020; Ullah et al., 2021; Khan et al., 2022; Ullah et al., 2022). Therefore, the present study focuses on the interactions between four non-native species communities with the hypothesis that hypothesized that the local environmental conditions associated with each community would provide a unique filter evident in the community floristic structure and composition of the resident species along with the disturbance and increases or decreases of diversity. In addition, the variation in communities across the spatial and soil variables was also questioned. The four non-native communities, i.e., Datura innoxia (Mill.), Xanthium strumarium L., Parthenium hysterophorus L., and Silybum marianum L. were focused in the study because these plant species rapidly spread in the region severely affecting the native communities’ diversity. The study’s aims were 1) to assess four non-native species communities and their characteristic floristic variations, 2) to assess the environmental variables affecting these communities and evaluate diversity and similarity indices, and 3) to evaluate essential factors in maintaining four non-native species communities.

In 165 sample sites, 107 species were identified, each with unique floristic features ( Figure 2 ). Annual and biannual species dominated communities, each contributing 50% (54) and 49% (52) based on life cycle distribution. Similarly, native and invasive species were extensively dispersed in the communities, i.e., native (52%) and invasive (41%). In contrast, herbaceous life forms comprised 75% of the species in the communities, while the rest were shrubs and trees. Among the four communities, the species’ habitat-wise distributions included wasteland (33%), arable land (32%), roadside (23%), and dry slop (11%). Most of the plant species found in the four communities belonged to Asteraceae, Fabaceae, Lamiaceae, Solanaceae, and Amaranthaceae ( S Table 1 ). The floristic comparisons of the four non-native communities are presented in Table 2 . The highest variations in communities’ comparison were presented by the status of the communities having P< 0.001 for χ² statistics. The associated plant invasive species were dominated in C-IV, i.e., 59% (13 species), followed by C-II, having 44% weightage. In contrast, the C-I and C-III were dominated by native species having 59% and 60% weightage of the community species. In life forms, the four non-native communities were dominated by herbaceous species having significant variation with χ² value of 24.49 and P< 0.05. Similarly, the communities show significant differences based on the life cycle duration, having a χ² value of 7.02 and P< 0.05. The annual species dominated C-I, II, and IV, while perennial species dominated C-III. Based on the C-I, II, and III habitats, the species was primarily found in roadside areas, followed by wasteland and arable land. However, in C-III, most species were inhabited by arable land and wasteland.

The non-native species communities were categorized into four ecologically valuable groupings (C-I, C-II, C-III, and C-IV) based on dominant species in terms of IVI ( Table 3 ). Notably, the second-most dominant species in each of the four communities differ uniquely, except for C-I and II, which are comparable but somewhat different in IVI, i.e., 5.72 ± 0.34 and 7.91 ± 0.88, respectively, for P. hysterophorus. Moreover, D. innoxia, with a codominance of C. sativa (17.28%) and X. strumarium (5.47%), characterizes the floristic assemblage in C-I. However, invasive P. hysterophorus (5.27%) and Chenopodium album (3.45%) also contributed to this community. The C-II was dominated by invasive X. strumarium (45.33%), having co-dominant species, i.e., D. innoxia (5.14%) and P. hysterophorus (7.91%), as well as numerous additional new species from their previous and following communities (C-I, III, and IV). P. hysterophorus (46.77%) dominates the C-III based on IVI dominance ( Table 3 ). Apart from C. dactylon (16.72%), none of the species had IVI higher than 5% in this community. Interestingly, the group IV vegetation consisted of S. marianum (76.30%) with S. nigrum (1.75%) and A. sativa (1.81%).

The four non-native communities were further categorized using Ward’s agglomerative cluster into groups ( Figure 3 ). C-I consists of three groups having P. hysterophorus co-dominant species of group I. In contrast, invasive species might be categorized as part of group II, i.e., the Datura-Parthenium-Cannabis community. The co-dominant species of group III was Medicago denticulate, also known as the Datura-Medicago group. This native species shows that group III is more stable than groups I and II. The C-II, X. strumarium invaded regions were divided into four ecologically significant groups based on different ecosystems and species composition. I was distinguished by the presence of Xanthium-Cannabis, with C. album and C. dactylon predominating. The overwhelming prevalence of invasive P. hysterophorus and C. album in group II set it apart. In contrast to group IV, this includes X. strumarium-C. sativa, group III has X. strumarium-D. innoxia. The co-dominant specie between groups I and IV was similar, but since additional distinctive species are present, they are considered separate groups. Three separate groups comprise the C-III of P. hysterophorus, with group-I being dominated by P. hysterophorus and two significant related species, namely C. sativa and C. dactylon. Similar to group I, groups II and III likewise have C. dactylon as the co-dominant species. In C-IV, most stands had a pure population of S. marianum, which was grouped as group-I in the cluster dendrogram. Nevertheless, three co-dominant species, namely Phalaris minor, Stylosanthes humilis, and Fumaria indica, were in the heavily invaded group (group II). Melilotus indicus, Solanum nigrum, and Euphorbia helioscopia were the three co-dominant species in partially invaded communities, i.e., group III.

In the floristics spectrum, 107 plant species were recorded in the studied region, belonging to 25 families. The Asteraceae, Amaranthaceae, Solanaceae, Brassicaceae, and Fabaceae had the most species diversity, while the other families had the least. This relationship between plant families and invading regions having low species richness has been reported globally. For instance, (Qureshi et al., 2014; Seifu et al., 2017; Iqbal et al., 2020) found that the Asteraceae, Poaceae, Amranthaceae, and Fabaceae were the dominant families of non-native species-dominated communities. Even though our data concurred with these findings, differences in taxonomic ranks at the family level may be related to geography and climate (Kosanic et al., 2018).

In the floristics composition, some of the species reported were of conservation concern, e.g., Calotropis procera, Artemisia scoparia, and Indigofera gerardiana were recently reported to be rare species in the region (Ilyas, 2015; Faisal and Khan, 2022). Native species become more susceptible to extinction due to non-native species disrupting the variety and composition of native plant ecosystems (Zimmermann, 1991). The ability of non-native species to alter the physical and chemical composition of local soils, in turn altering community diversity, is an essential factor that assesses its potential to become naturalized (Iqbal et al., 2021). The communities were dominated by annual and perennial plant species, which is related to the opposing theories put forth in the literature that 1) annual plants, in particular, have significant effects on the homogenization of communities (Qian and Guo, 2010) and 2) perennial plants encourage community disturbance and invasion (Milchunas et al., 1988; Tappeiner et al., 1991). However, given the propensity for bigger plants to displace smaller ones over time, little herbs may be disadvantaged (Timsina et al., 2011). In areas where herbs or shrubs predominate, invasion of certain plant species was dominant on the sides of highways and waste fields, as well as in agricultural fields. Similar predictions have been made by (Burleigh, 2005; Brisson et al., 2010; Richardson, 2011; Davies et al., 2013; Salgado et al., 2019) that agricultural fields, railroad lines, and road pavements would be the most vulnerable to plant invasion.

Low diversity indices were recorded in the study region, possibly related to the high invasion rate at the assessed sites. Several researchers, e.g., (Seifu et al., 2017; Qureshi et al., 2019; Iqbal et al., 2020; Iqbal et al., 2021), reported that invaded sites have low diversity indices. In non-native populations, Shannon diversity and IVI showed a declining tendency along invasion gradients. For instance, colonies of S. marianum that have been entirely or severely invaded have low diversity indices. In invaded areas, Seifu et al. (2017); Qureshi et al. (2019), and Iqbal et al. (2020) demonstrated poor diversity of vegetation in sites invaded by non-native species. However, the number of species in a community is little impacted by non-native naturalized species (Awasthi et al., 2012), such as in the communities of D. innoxia. Similarly, D. innoxia and Cannabis sativa are native invasive herbaceous plants with high growth rates, which may encourage a rise in native, fast-growing species due to compositional changes.

In environmental variables, elevation was a prominent factor in every community, i.e., C-I-IV, across which the vegetation characteristics vary ( S Table 2-5 ). The species were most abundant at lower elevations, as high elevations have more diurnal variation in the climate (Zia et al., 2018). Along the elevation gradient in non-native communities, our results showed distinct and substantial variations in community composition across environmental variables. For instance, as the altitude rises, the percentages of OM, N, and available water decrease while the percentages of lime, silt, and lime increase. Elevation impacts soil characteristics (Blackburn et al., 2011), while (Kosanic et al., 2018) reported that a plant species’ capacity to invade the soil indirectly or directly impacts the chemical composition and availability of nutrients. According to research, invasive and naturalized species may use the nutrients in the soil to control their establishment and reproduction (Qian and Guo, 2010; Kosanic et al., 2018). The consistency of soil nutrients may also result from alien naturalized/invasive plants, which might encourage future invasions into populated regions (Qian and Guo, 2010). Carbon, nitrogen, phosphorous, and potassium increased in invaded areas (Bekele, 2000). In nutrient-rich situations, plant growth rate and biomass production increased concurrently with changes in plant biomass (Tappeiner et al., 1991), consistent with our findings. Environmental variable alterations have been shown to assist invaders like X. strumarium (Gallardo et al., 2017) and P. hysterophorus disruption (Milchunas et al., 1988), and this may also be true for D. innoxia. A combination of soil instability and fertilizer inputs is crucial for promoting and establishing non-native plants (Salgado et al., 2019). The relationship between plant invasion and soil nutrition is well established (Davies et al., 2013) and is shown by the increased degradation and microbiological activity in P. hysterophorus-infected soils (Brisson et al., 2010). D. innoxia has experienced the same enhancement of microbiological activities as evident from nutrient-rich zones that were primarily invaded. Based on our findings, invasion and naturalization intensities may be restricted at higher elevations, as evident from decreased non-native species IVI in higher elevation sites.

Although the precise mechanism and year of the community invasion in the area are unknown, non-native species have unintentionally been introduced into several Khyber Pakhtunkhwa regions over the past two decades during the movements of Afghan refugees, flood runoff, and tourist visits for recreational purposes (Gaertner et al., 2016; Gallardo et al., 2017). These non-native species have been widely established, naturalized, and multiplied, negatively affecting biodiversity and agriculture (Marwat et al., 2010). Depending on how well they adapt to the new environment, non-native plant species in new habitats may either grow or perish (Ullah et al., 2022). It is challenging to foresee whether these would flourish and naturalize in the new setting (Sherley, 2000; Marwat et al., 2010; Gallien and Carboni, 2017). The successful, invasive plants often spread, altering vegetation and environmental factors (Zimmermann, 1991; Le Maitre et al., 2011; Ratnayake, 2014; Ullah et al., 2022) and endangering native plants and animals (Zimmermann, 1991; Qureshi et al., 2019). Our research confirmed prior assertions that these invasive non-native plants have effectively supplanted local plant species, colonized the semi-arid parts of Khyber Pakhtunkhwa, Pakistan, and established themselves as significant invasive species there.

The findings of the research suggest that the area is invaded by different types of non-native species, which may be naturalized (D. innoxia), invasive (X. strumarium), and exotic (P. hysterophorus and S. marianum). In these species, S. marianum form pure communities having homogenized communities. In contrast, the P. hysterophorus communities were diverse, consisting of various native and non-native species suggesting that invasive species can decrease or increase biodiversity. The research indicated that the communities in the area are at risk, and proper strategies are needed to control and overcome the spread of these non-native species. By conducting qualitative and quantitative monitoring using species inventories (seasonally) and phytosociological methodologies, as well as mapping using ground-based techniques (through map overlays or GPS) and remotely sensed images (aerial photos, high-resolution multispectral digital data), future invasions of such invasive species may be prevented. Researchers need to engage with taxonomists, ecologists, and land managers to establish a plant detection network in vulnerable areas to prevent such species from becoming invasive. The CCA analysis of each community revealed the importance of elevation gradient in spreading these invasive species. Therefore, the environmental gradient must be considered while devising strategies for controlling these invasive species.

Moreover, the life cycle of these species in most cases was also non-overlapped, indicating the disturbance of communities in the whole growing period. In addition, there is also a need to highlight the use of these invasive species in ecosystem services as bioresources for different purposes. The biomass of invasive species may also be utilized for medicinal and phytochemical purposes, and some of these species, such as X. strumarium, have recently been reported for heavy metal pollution biomonitoring and phytoremediation (Ulalh and Khan, 2022). We also advise assessing these species’ morphological, phenological, phytochemical, physiological, and reproductive biology for efficient management. In addition, as suggested by Dickie et al. (2014), we argue to assess the positive and negative traits of such species be investigated in a broad, systematic framework to prevent any inconsistencies between their risks and benefits.

The original contributions presented in the study are included in the article/ Supplementary Material . Further inquiries can be directed to the corresponding author.

NK, and RU conceptualization this research, conducted the field and laboratory experiments, and compiled and analyzed the data; NK and RU wrote the original draft of the manuscript; MO, and MA-M supervised the experiments; MO, and MA-M, contributed to data analysis, helped in review and edited the initial draft of the manuscript; all authors read and approved the manuscript.