literature review on species composition

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• Published: 01 July 2020

Historic changes in species composition for a globally unique bird community

• Swen C. Renner   ORCID: orcid.org/0000-0002-6893-4219 1 , 2 &
• Paul J. J. Bates   ORCID: orcid.org/0000-0003-3630-739X 2 , 3  

Scientific Reports volume  10 , Article number:  10739 ( 2020 ) Cite this article

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• Biodiversity
• Community ecology
• Conservation biology
• Environmental impact
• Tropical ecology

Significant uncertainties remain of how global change impacts on species richness, relative abundance and species composition. Recently, a discussion emerged on the importance of detecting and understanding long-term fluctuations in species composition and relative abundance and whether deterministic or non-deterministic factors can explain any temporal change. However, currently, one of the main impediments to providing answers to these questions is the relatively short time series of species diversity datasets. Many datasets are limited to 2 years and it is rare for a few decades of data to be available. In addition, long-term data typically has standardization issues from the past and/or the methods are not comparable. We address several of these uncertainties by investigating bird diversity in a globally important mountain ecosystem of the Hkakabo Razi Landscape in northern Myanmar. The study compares bird communities in two periods (pre-1940: 1900–1939 vs. post-2000: 2001–2006). Land-cover classes have been included to provide understanding of their potential role as drivers. While species richness did not change, species composition and relative abundance differed, indicating a significant species turn over and hence temporal change. Only 19.2% of bird species occurred during both periods. Land-cover model predictors explained part of the species richness variability but not relative abundance nor species composition changes. The temporal change is likely caused by minimal methodological differences and partially by land-cover.

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Introduction.

Species richness, relative abundance and species composition are dynamic phenomena and vary in space and over time 1 , 2 . Recorded fluctuations of bird species richness and species diversity are explained by deterministic changes (e.g., global change such as changes in land-cover or land-use intensification 1 ), methodological changes (different effort or sites; limited or no standardisation; and methodological and non-systematic errors 3 , 4 ), random processes (e.g., neutral dynamics 5 , 6 , 7 , 8 ), or any combination of the above.

Global change is rapidly proceeding and includes land-use intensification, changes in land-cover, climate, atmospheric composition, and invasive species, among other factors 1 . Land-cover change is probably the most important in terms of species response 9 , 10 . It is probably more important than climate change for most biodiversity 1 , 11 , particularly for many bird species 9 . Land-cover change, including land-use intensification, have been shown to affect species in a variety of direct, indirect and interacting ways, including local extinction, range shifts, changes in local abundance, or interactions with other species 12 , 13 , 14 .

Although several studies have shown the effects of global change in the form of habitat loss or land-use change, these studies typically are limited in explanatory power. In many cases, the historic (previous) baseline, which is used to estimate the diversity statistics, has a low statistical power. In others, the temporal aspect is too short to show meaningful effects 15 . Most studies use only recent baseline data and the time difference (mostly 2 years, in rare instances more than a decade) is too short for changes in species assemblage. Typically they explain only short-term fluctuations, particularly fluctuations within or between consecutive years 7 , 16 , 17 .

Only a few regions worldwide remain with a habitat cover of near pristine condition 15 . These untouched areas are embedded in a land-cover mosaic of various forms 18 . The Hkakabo Razi Landscape in the northern tip of Myanmar is largely untouched and includes large tracts (11,280 km 2 ) of pristine forests interspersed with a few, relatively small areas of degraded forests or other local land-cover forms 19 . Within this Landscape, an historic bird assemblage has been documented 20 . These baseline data were collected by British and US explorers and include specimens and letters on methods and localities. It is the quality of this historical documentation, together with the rigour of the collecting methodology (which compare favourably and complement the recent efforts), that makes the Hkakabo Razi Landscape almost unique for studying species compositional turnover. At the same time, the effects of land-cover change on birds can be analysed, because the historic samples are located in pristine forests, while our own recent samples cover both pristine and some of the relatively few degraded habitats. The historic and recent collections, when reviewed together, allow for an analysis of the historical changes to bird assemblages, covering data separated by almost 70 years. This will increase our understanding of temporal dynamics in bird communities.

Mountain ecosystems of the tropics are home to high species diversity. Those of the Himalayas, including the Hkakabo Razi Landscape are also home to a rich variety of endemic taxa. For example, the study area hosts at least one endemic bird ( Jabouilleia naungmungensis 21 ) and at least two endemic subspecies of birds 20 , 22 ( Alcippe cinereiceps hkakaboraziensis , Malacocincla abbotti kachinensis ; all species and samples are listed in Online Supporting Information Table S1 ). However, all biodiversity in mountain ecosystems is vulnerable to land-cover change 23 . Currently, the forests of the Hkakabo Razi Landscape are likely the last vast area of pristine forests in Asia or at least Southeast Asia 19 , 24 , 25 , 26 , with relatively few degraded habitats imbedded within the pristine forests. To date, 456 bird species have been recorded in the Hkakabo Razi Landscape, proving its global importance for bird conservation. While the Hkakabo Razi Landscape covers about 1% of terrestrial Myanmar, it is habitat for almost half of all bird species recorded from the country (456 vs. ~ 1,100 20 ).

Here we describe and test species turnover and temporal variation in relation to global change parameters. We predict no detectable differences in species richness, relative abundance or species composition between the two periods considered (pre-1940 vs. post-2000), because land-cover change has not yet occurred to a significant extent. In turn, any significant differences in species richness, relative abundance or species composition would indicate a high proportion of temporal variation (i.e. non-static species composition) and/or a response to deterministic reasons (e.g., environmental drivers). If we find a differences in species composition between the two periods of over 50% (natural fluctuations in bird communities of the tropics with very limited human impact exhibit up to 49% species change within or between years 27 ), or significant variation between the periods in respect to species richness and relative abundance, these fluctuations can be interpreted as a consequence of temporal change.

Since the Hkakabo Razi Landscape is one of the few remaining significantly large and natural mountain forests worldwide 24 , 25 , 26 , 28 , from which we have almost perfect historical datasets, it is an invaluable natural laboratory in which to test the impact of temporal change on species richness, relative abundance and species composition. The results of such studies are of global importance. The Hkakabo Razi Landscape is a unique constellation of largely “untouched” forests 26 with collectors in the first half of the twentieth century having labelled their specimens almost perfectly.

Material and methods

Study region and study sites.

The study sites, i.e. the localities of bird sampling, are located in the Hkakabo Razi Landscape. Distribution of the localities and consequently the area covered is defined by the historic collectors (redrawn in Fig.  1 , following Suarez-Rubio, et al. 26 ). The Hkakabo Razi Landscape is located in the northern most part of Myanmar (to many Westerners still known as “Burma”) and comprises the Hkakabo Razi National Park, the planned “Southern Extension” of the National Park and the Hponkan Razi Wildlife Sanctuary (all borders as proposed on August 15, 2015).

Map of study region in northern Kachin State, Myanmar ( red area in inset map shows the location of the protected areas and outlines the Hkakabo Razi Landscape within Myanmar). Blue and green circles are for sample sites from which data are used in the study. Grey and open circles are for sample sites whose data are excluded since they are either outside the study region or have incompatible datasets.

Within the Hkakabo Razi Landscape, two major survey programmes have been completed to assess the total number of species. The first, a series of uncoordinated ‘one-off’ studies was undertaken by British collectors in the early twentieth century, the second by S.C.R. and several colleagues in the early 2000s 20 . All samples are available, either at the Natural History Museum (Tring, UK), or at the Smithsonian Institution (Washington, DC, USA) or at the Zoological Park (Yangon, Myanmar).

In the post-2000 studies, all samples were taken in accordance with European Union, US and particularly national Myanmar laws on animal protection and conservation measures at the time of data sampling. All necessary permits have been approved by the Nature and Wildlife Conservation Division of Myanmar’s Ministry of Natural Resources and Environmental Protection (MoNREC, formerly Ministry of Forestry—MoF). The responsible officers and the permit number are listed in the acknowledgements.

Collection based data search

There is an enormous amount of data available from bird collections worldwide. However, insufficient or imprecise locality data and habitat information is an issue for analysis involving museum specimens, particularly if collected prior to the 1960’s. Nevertheless, a quite remarkable number of specimens in the collections is available for the Hkakabo Razi Landscape for further analysis. Those collected by the British forester Ronald Kaulback (sometimes written as Kaulbach) and his colleagues indicate on the labels exact locality, including coordinates and elevation (details listed in Table S1 , Online Supporting Information). They also provide simple information about the habitat types and how the birds were captured. Many were collected in the Adung Valley, which is today part of the Hkakabo Razi National Park. Kaulback was loosely associated with Lord Cranbrook, Francis Kingdon-Ward, and Bertram C. Smythies 29 , 30 , 31 , 32 , 33 , the latter used much of Hebert Cecil Smith’s information from “Notes of the birds of Burma” 34 in his field guide “Birds of Burma”. Smythies 31 and Mayr 35 provided detailed sight records of the birds found by Kaulback, and by Major John K. Stanford 36 , 37 , 38 , 39 , 40 , 41 , 42 , 43 , 44 , 45 , 46 , 47 , 48 , 49 , 50 and Garthwaite 34 . In addition, the botanist Francis Kingdon-Ward 51 , 52 , 53 , 54 , 55 , 56 , 57 provided some additional specimens to the collections (Kaulback participated in some of Kingdon-Ward’s expeditions). The specimens used for this analysis, were collected by Kaulback (96 specimens), Lord Cranbrook (34), Stanford (30), Kingdon-Ward (12), and an anonymous collector (possibly identified as Kingdon-Ward based on the collection date and locality: 1). All these records are the baseline for reconstruction of the historic bird community used herein.

All the historic specimens are held in the Ornithology Section of the BMNH in Tring, UK (collections visited during the study are listed in “Appendix A”, Online Supporting Information). They were collected between January 1, 1900 and September 20, 1939 (16 specimens are labelled without any date). However, most of the specimens were collected in 1931 and 1938, while the other years show a relatively low coverage (Fig.  2 ).

Individuals (specimens) sampled per year in the Hkakabo Razi Landscape pre-1940 and post-2000 with comparable methods and effort.

Kaulback and colleagues used mainly shotguns in the Hkakabo Razi area and also set snare traps 58 . Kaulback gives a rough indication of his shotgun collecting effort. It is apparent that there was probably two individuals shooting birds for a maximum of “half an hour” each day 58 . The effort for the snare trapping is not documented. However, from the labels written at the time of collection, ~ 5% of the specimens in the Hkakabo Razi Landscape were described as “snared in”. The specimens were collected on a total of 56 capture dates for pre-1940 (dates derived from labels in the collections). This is likely to equate to the number of days Kaulback and his colleagues sampled birds with shotguns or snare traps.

The historic collection covers an area between 27.10 North degree to 28.50 North and 96.50 East to 98.40 East (Fig.  1 ). The historic spatial extent has been chosen with a maximum overlap with the recent collection in order to maximise the comparison between the two periods. Therefore, we neglect in any analysis further localities of historic and recent collectors, particularly towards the South of the study sites (the so-called “The Triangle”) and east in Yunnan (China) (Fig.  1 ). The historic sites used in this study, sum up to 17 and have been sampled mainly in February/March (13% of the specimens pre-1940), July–September (34%), November-January (31%) in 1931 and 1938. We verified all data to the best of our knowledge to maximize accuracy and precision.

Only one historic record has been corrected based on inconclusive data (post-museum procedure), because “Adung valley” is certainly not at 97.00 East (as written on specimen label), because 97.00 East is located in India (BMNH 1938.5.5.1, Strix aluco , Female, collected by Kingdon-Ward on 7 March 1937). We corrected the locality’s coordinates in our database by using the coordinates as from other specimens’ labels with the same locality name (“Adung Valley”).

Recent collection and field data accumulation

The post-2000 data were collected between February 9, 2001 and March 20, 2006 (Fig.  1 ). This dataset comprises collections made by John H. Rappole and S.C.R., with significant support from Thein Aung, Nay Myo Shwe, Myint Kyaw, Myint Aung, and Chris M. Milensky 19 , 20 , 21 , 59 , 60 , 61 , 62 , 63 . For all recent sampling included in this study, mist nets of 12 m × 2.6 m have been used (number of capture days and mist nets detailed in Table 1 ). Typically, nets were set from 05:00 to 10:00 and from 15:00 to 18:00 local time. Recent sites were sampled for 2 days in 2001 and for 1 day each in 2004, 2005, and 2006 (details on capture days and net numbers in Table 1 ). The sampling season was mainly February to March for the recent sites and there are 65 capture days for the post-2000 period included in this analysis (Table 1 ).

The methods during historic (snares/shotguns) and current data collections (mist nets) imply differences employed in collection methods. Consequently, some difference based on the methods might explain at least part of the differing species composition (further details in the discussion). However, the datasets are very close—if not identical—for several characteristics: elevational band (between 400 and 2000 m in both periods, with an occasional locality from a higher elevation); capture localities within a small spatial margin (maximum distance between the sites is 25 km; Fig.  1 ); days of capture (56 vs. 65). The number of sites within the general study area was 17 versus 17. However, the exact localities are different (Fig.  1 ). The selection criteria for the historic sites is not documented. The recent sites were chosen randomly and are within walking distance of an existing settlement. Historic and recent sites were selected independently from each other.

We use three terms to characterize the bird communities (which are also the dependent variables in the tests as outlined below). Species richness is the number of species within the collections. We use the number of species per period (or per land-cover type or per locality) to have a measurement for comparison. With limitations, it is also possible to establish a measurement of abundance for each of the two periods, since all collections show for several species different numbers of sampled individuals per capture site (i.e. the number of samples detected for each species per site). This is described here as relative abundance . The number of individuals sampled per species and period is a measurement of abundance. The species composition comprises a list of species names per site (or locality or period).

Land-cover and habitat data

For Hkakabo Razi Landscape, land-cover and land-use classifications are available from 1989 26 , 2001 19 , and 2016 26 . Land-cover has been classified as mostly intact. The changes in land-cover types are marginal with an annual deforestation rate of ≤ 0.23% from 1989 to 2016 26 (this is indistinguishable from background noise). There has been almost no change in land-cover for the historic bird localities from 1989 to 2016. Only 47 times the land-cover form of the two classifications changed from 1989 to 2016 (Table 2 ) for the exact localities of the 173 recorded pre-1940 birds (the Landsat 30 m by 30 m pixel 26 ). Most of these land-cover changes are negligible and based on melting snow (e.g., from snow/ice/glacier to rock/bolder). Except for the Putao plains, hardly any change occurred from 1989 to 2016 19 , 26 . The pre-1940 localities have remained undeveloped 58 , 64 , 65 or are close to the same settlements as in 2016. Since most of the habitat remains pristine, all historic (pre-1940) localities are assumed to be of the same land-cover type as classified in 1989. All post-2000 localities are assigned to the land-cover classification as of 2016.

The land-cover has been classified with the same class names for 1989 and 2016 26 : pine/rhododendron, forest < 600 m, forest 600–1800 m, forest > 1,800 m, grassland/pasture, ice/glacier, clear-cut, paddy field, rock/boulder, secondary forest < 600 m, secondary forest 600–1800 m, settlement, shrub/bush/fern, and streambed.

Statistical outline

The largest issue for testing is the unknown sampling effort for pre-1940. It is not possible to assess the completeness of the pre-1940 datasets because this information is not available in the archives or from the specimen labels. This makes for some uncertainty when comparing the bird assemblages from the two periods. Nevertheless, we maximized comparability of the datasets, as far as possible with the archival and label information available. Details on methods are further published for post-2000 20 , 59 .

We performed an ANOVA for species richness and relative abundance to analyse the variance between the periods (pre-1940 vs. post-2000) and used a Kruskal–Wallis test if not normally distributed. In a second step, we added a generalized linear model (GLM) approach to test whether habitat or locality (sampling sites pooled as per nearest settlement name) have an effect on the species richness or relative abundance. Differences between periods and species richness (or relative abundance) were assessed using analysis of variance (ANOVA) after verifying for homogeneity of variances (Fligner test) and normality (Bartlett test). All analyses were performed in R version 3.5.1 66 and an α-level of 0.05. Observational data (count, species numbers) have been log-transformed (ln).

We assessed differences in bird species composition among periods, habitat types and localities using non-metric multidimensional scaling (NMS). Relative abundance was square-root transformed ( vegan -package). The NMS was run using Sørensen (Bray–Curtis) distance with an automatic stepping down resolution starting 200 runs from a random configuration.

Since the number of sample sites and the number of samples overall is relatively low for all analyses, we performed power analysis with the pwr -package in R to assess the strength of statistical outcomes. We assessed power always with a significance level of 0.05. ANOVA (variation of species numbers between periods) had a power of 1; GLM with response “species numbers” 0.137 and power for GLM with response “relative abundance” was 1.

There were a total of 708 individual bird records belonging to 193 species; 98 species (173 individuals) were only recorded pre-1940 and 132 species (535 individuals) were only recorded post-2000. Only 19.2% (37 species) occurred in both periods. This indicates a considerable discrepancy between the species assemblages. The top five most abundant species differed between the periods: Post-2000, the most abundant species was Alcippe morrisonia (42), followed by Alcippe rufogularis (25), Alophoixus flaveolus (19), Niltava grandis (19), and Ficedula monileger (18). For pre-1940, the most abundant species was Garrulax striatus (7), followed by Aethopyga saturata (6), Heterophasia pulchella (6), Arachnothera magna (4), and Cissa chinensis (4) (all samples and species included are listed in Table S1 , Online Supporting Information).

Analysing species composition with NMS yielded weak ties and hence should be considered with caution. Nevertheless, model-selection procedures in NMS showed that “period” is the best explaining factor out of “period”, “habitat (2016)” and “locality” (CCA stepwise permutation selection p for “period” = 0.02, all other p  > 0.05). Contrasting to the species composition, species richness showed no differences. Species richness did not change from pre-1940 to post-2000 (Kruskal–Wallis χ 2  = 3.774, df  = 1, p  = 0.052; Fig.  3 ). When modelling species richness with the predictors “locality”, “period”, “land-cover 1989” and “land-cover 2016”, the latter two had an effect on the species richness ( p  < 0.001, GLM models “s2” and “s3” in Table 3 ). Considering each predictor singly with species richness, only “period” predicts species richness (models “s4” to “s7” in Table 3 ).

Species numbers ( A ) and relative abundance ( B ) per collecting locality of birds in the Hkakabo Razi Landscape pre-1940 and post-2000 with comparable methods and effort. The black solid line indicates the median, circles indicate outliers, whiskers 95% CI and box margins 75% CI.

The relative abundance changed from pre-1940 to post-2000 (Kruskal–Wallis χ 2  = 26.125, df  = 1, p  ≤ 0.001; Fig.  3 ). When modelling relative abundance, neither “land-cover 1989”, “land-cover 2016”, nor “period” had an effect on the relative abundance, but only “locality” ( p  < 0.001, GLM model “a1” in Table 3 ). Considering each predictor singly with relative abundance, all but “species” predict relative abundance (models “a3” to “a7” in Table 3 ).

The community structure follows a typical species rank-abundance curve (Fig.  4 ), with few species of many individuals and many species with few individuals. The species-abundance shows a similar pattern for pre-1940 and post-2000, however, the post-2000 is about one magnitude higher (Fig.  4 ).

Species rank-abundance (i.e. number of specimens) curve of all detected individuals per period considered in the analysis of birds in the Hkakabo Razi Landscape pre-1940 and post-2000 with comparable methods and effort.

The long-term datasets of Hkakabo Razi Landscape inform us that while species richness did not change from pre-1940 to post-2000, species composition and the relative abundance changed significantly, including the five most abundant species in each period. The magnitude of change in species composition, with less than 20% of taxa shared between the two time periods, is particularly noteworthy and unexpected.

In a world where temporal patterns of biodiversity have received much less attention than spatial ones 67 , 68 , the datasets from Hkakabo Razi Landscape are important because, almost uniquely, they give us the chance to differentiate between anthropogenic impacts and background temporal changes in ecological communities in an extensive area of Old World forest biome, with a timescale of almost a century (primarily between 1931 and 2006, although a minority of specimens were collected as far back as 1900). The datasets are unusual for such studies because they are based in a subtropical rather than a temperate area and are drawn from a large tract of forest (11,280 km 2 ) that remains almost pristine. Furthermore, they are statistically valuable since, although the methodologies between the historical and more recent surveys are not the same, they are well documented and share many comparable components and are sufficiently informative to give us the opportunity to observe temporal changes not only in species diversity and species richness, but also crucially in species composition, and to a lesser extent, relative abundance within species.

The results from Hkakabo Razi Landscape, particularly the large variation in species composition, reflect earlier findings from Costa Rica, where working with more detailed data, albeit gathered over a much shorter time-scale (1985–1992) 67 , researchers noted that tropical bird communities far from being stable systems are in reality dynamic ones with a ‘complex mix of stable and variable components that produce changes in species composition and abundance over various spatial and temporal scales’. This variability in Costa Rica was observed not only, as might be expected, in secondary forest (partially as a response to vegetational succession) but also, though to a lesser extent, in mature forest. It should also be noted that rates of temporal turnover will also vary amongst ecosystem types 67 , 68 and in relation to local environmental factors, with variable responses to the same disturbance events 68 .

In contrast to the Hkakabo Razi Landscape study, which provides information on long-term temporal patterns, most others have focused more on short-term fluctuations driven by resource availability 69 , 70 . These include, for example, the movement of birds in response to the availability of fruits in a mountain biome in Costa Rica; the differential movement of insectivorous and frugivorous birds in Kenya in response to food availability; and the movement of birds in the Australian tropical forest in response to climatic variations and subsequent resource availability 71 . Other studies, both short and long-term, and over a variety of spatial scales, have focused on changes in bird diversity and composition but primarily in areas that have been significantly impacted by anthropogenic activities. These include, for example, studies of the temporal variation of taxonomic and functional diversity in the conterminous USA based on 40 years of data (1970–2011) 72 , 73 , 74 . Such studies, although extremely valuable, do not provide data that enables us to develop conservation policies that take into account purely natural cycles in diversity and abundance.

Without an understanding of natural long-term variability in essentially pristine ecosystems, it is almost impossible to differentiate between human-induced change and natural cycles in those that are anthropogenically modified. As such, the impacts of human induced environmental change may be overstated when comparing differences in species composition at any particular site over a longer time period.

That said, the results of Hkakabo Razi Landscape, should be treated with some caution. For although some of the variables in the collection methods between the pre-1940 and post 2000 datasets are (surprisingly) comparable, others are not. Those that are similar (as outlined in the Methods section) include: elevational band, primarily between 400 and 2000 m in both periods; number of collection sites, 17 versus 17; spatial distribution of capture localities, which although not the same, have a maximum distance between the sites of 25 km (Fig.  1 ); the number of capture days (56 vs. 65); period of collection 8 years (primarily from 1931 to 1938) and 6 years (from 2001 to 2006)—all of this is important since typically it has been predicted from elsewhere that there will be around twice as many species detected in a decade as in a single year 68 . However, there are also differences, the most important of which is capture method. This could be particularly important in an ecosystem, where it is predicted (based from data collected elsewhere) that high species diversity is inversely correlated to low species density—i.e. many species with fewer individuals. It is probable that some of the difference in species composition observed from the pre-1940 post 2000 data is directly attributable to differences in collecting method. Post-2000, the exclusive use of mist-nets would favour the collection of those bird species that favour niches nearer to ground level, whilst pre-1940, a hunter with a gun, will have greater success with birds, which are more visible and/or high in the canopy. This is reflected in the five most abundant species of pre-1940, which are either more colourful (e.g., brightly coloured such as some laughingtrushes), or easy to watch (such as Arachnothera magna which occurs in open forest patches and at the forest edge), or more visible through their behaviour (e.g., loud alarm calls such as from Garrulax striatus ). Contrasting, the top five post-2000 species are more secretive in behaviour and less bright coloured, hence less obvious to the hunter.

Furthermore, the by one magnitude higher, relative abundance post-2000 is probably a methodological bias. While mist nets capture, for example, the largest part of an Alcippe morrisonia flock (20 + individuals, own unpublished observations), the hunters pre-1940 shot one individual out of a flock, and the remainder of the flock certainly escaped and disappeared without trace in the forest. Moreover, mist nets, unlike hunters, do not discriminate since they catch every bird that becomes entangled in them whereas a hunter may either consciously or subconsciously eschew birds of a species for which a number of specimens have already been collected. Theoretically, the only way to compare the relative abundance between the two periods would be to collect birds today in a manner similar to that employed pre-1940. However, these methods, shooting and snares, are obviously not possible or desirable today for ethical reasons and Myanmar national laws.

In addition to variation in capture methods, there is some variation in the season of capture between the two datasets. Post-2000, all 535 individuals of the 132 species were collected in the months February–March. However, for the 173 specimens collected from pre-1940, 13% were captured in the February–March time period whilst the remainder were collected mainly in July–September and November-January. This is important since Myanmar hosts a diverse winter migrant bird fauna and since inter-seasonal fluctuations in bird composition are known to be on average higher for migratory and nomadic species than for sedentary ones 70 , 75 , 76 . However, interestingly, hardly any long-distance migrants were detected in either the pre-1940 or post-2000 datasets so that migration status alone cannot explain the large fluctuations seen in species composition between the two time periods.

An additional analysis including, for instance, the phylogenetic structure 77 of the bird community or its functional traits, could add further insights. However, for this paper we have avoided such approaches since, currently, the phylogenetic structure of the phylogenetic placement and validity of the three most important families in our data set, the Muscicapidae, Timaliidae, and Pellorneidae, are controversial and all deep-phylogeny assignments are in continuous flow for many species occurring in the Hkakabo Razi Landscape (detailed in Online Supporting Information C). Meanwhile data on the functional traits of bird species from Hkakabo Razi Landscape remains incomplete and/or speculative with little detailed information on the functional groups beyond generalised descriptions, such as insectivores, granivores,… There are also no data available on seasonal variation, e.g. breeding versus non-breeding 18 . Therefore, rather than working with incomplete or speculative data sets, we focused on the parsimonious and relative robust analysis of the bird community.

Preliminary analysis of the long-term Hkakabo Razi Landscape datasets provide some very interesting information that is of importance not just to bird ecologists but to the much broader scientific community, especially those concerned with environmental change, including climate change and habitat fragmentation, and its impact on biodiversity. The datasets help put short-term fluctuations into a meaningful context, for example within monitoring programmes, and provide information that gives an insight into whether contemporary trends in diversity are simply a response to anthropogenic-induced changes or are the result of dynamics originating before the onset of the Anthropocene 78 , 79 . They also have important implications for conservationists who seek to interpret the meanings of changes in faunal composition both in natural and man-made habitats and who wish to develop conservation policies that take into account natural cycles in diversity and abundance. As with interesting studies in the USA and France, the next stage for the Hkakabo Razi Landscape data are to develop more sophisticated models to determine if significant changes in taxonomic diversity are also reflected in changes in phylogenetic and functional diversities 79 , 80 , as well as determining random portion of the species richness 5 , 6 , 7 , 8 .

The two Hkakabo Razi Landscape datasets, pre-1940 and post-2000, give an invaluable insight into the question ‘what is the underlying level of temporal turnover in a bird community?’ They help us to understand background turnover in birds in a subtropical pristine forest site, which will provide an invaluable foundation (despite the caveat of different methodologies in the two datasets) when trying to assess anthropogenic impacts in increasingly disturbed habitats elsewhere. The datasets further challenge the notion that bird communities in the tropics/subtropics, even in natural habitats, are stable systems. Rather they show that there is an important temporal component to biodiversity and that natural ecosystems are dynamic with a complex combination of stable and variable components and that this dynamic component impacts in different ways and with different severity on species diversity, species composition, and relative abundance.

Data availability

The data used for analysis is available in the Online Supporting Information.

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Acknowledgements

Our special thanks are for John H. Rappole, who invited S.C.R. the first time to join him on an expedition to the Hkakabo Razi National Park in 2001. We would also like to thank all friends and colleagues supporting the fieldwork of this study during the last two decades, namely: Marcela Suarez-Rubio, Thein Aung, Nay Myo Shwe, Myint Kyaw, Sang Nai Dee, Myint Aung, Braing Shaw, Kyaw Lin, Tu Myint U, A Jo, Chris Milensky, Tay Zah, Aung Maung, Aung Kyaw, Naing Lin, Dee Shin, Htin, Hdoa Dee and over 100 helpers during the eight trips to the region. We thank Aung Khin and Thandar Kyi, who organised the expeditions in 2004, 2005, and 2006, and Tay Za who managed the 2001 trip. We would like to thank the curators, managers and technicians of the various collections S.C.R. visited for the study, namely (indicating the collection while they were employed there during the visits): Martin Päckert SNHD, Paul Sweat AMNH, Leo Joseph and Nate Rice ANSP, Mark Adams and Robert Prys-Jones BMNH, Jack Dumbacher and Moe Flannery CAS, Sylke Frahnert MfN, Hein van Grown NMN, Carla Dove, Gary Graves, Helen James, and Terry Chesser NMNH, late Anita Gamauf NHMW, Ulf Johansson NRM, Freddy Woog SMNS, Jon Fjeldså ZMUC, Till Töpfer and late Stefanie Rick ZFMK. Funding to visit collections was provide by the European Union SYNTHESYS framework (FR-TAF-6275, DE-TAF-6206, ES-TAF-2501, AT-TAF-2481, GB-TAF-108, SE-TAF-1312, NL-TAF-4369, GB-TAF-4367, DK-TAF-4963), and to visit the field by the National Geographic Society (GEFNE48-12). Besides all professional disagreement, we thank Hannah Fraser plus two anonymous reviewers and the editor for their valuable input on previous versions of the manuscript. Late Uga, former Director of the Nature and Wildlife Conservation Division, initiated the 2001 trip and we thank him for his efforts to make the expeditions possible. Without his support, none of the work done in the region would have occurred. We thank the Nature and Wildlife Conservation Division of the Forestry Department, and especially former Director Khin Maung Zaw, for permission to conduct the study (Myanmar Collection and Export Permit # SI/4697/2004). Tin Tun was implementing the permits in 2005 and 2006. Open access funding provided by BOKU Vienna Open Access Publishing Fund.

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S.C.R. developed the study, collected that data, analysed all species’ statistics. P.J.J.B. scrutinized each sentence for English and supported significantly the discussion section and all discussions with Reviewer B. S.C.R. and P.J.J.B. equally wrote, contributed and finalized the text.

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Renner, S.C., Bates, P.J.J. Historic changes in species composition for a globally unique bird community. Sci Rep 10 , 10739 (2020). https://doi.org/10.1038/s41598-020-67400-z

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REVIEW article

Pressures on boreal riparian vegetation: a literature review.

• Department of Environmental and Life Sciences, Karlstad University, Karlstad, Sweden

Riparian zones are species-rich and functionally important ecotones that sustain physical, chemical and ecological balance of ecosystems. While scientific, governmental and public attention for riparian zones has increased over the past decades, knowledge on the effects of the majority of anthropogenic disturbances is still lacking. Given the increasing expansion and intensity of these disturbances, the need to understand simultaneously occurring pressures grows. We have conducted a literature review on the potential effects of anthropogenic pressures on boreal riparian zones and the main processes that shape their vegetation composition. We visualised the observed and potential consequences of flow regulation for hydropower generation, flow regulation through channelisation, the climate crisis, forestry, land use change and non-native species in a conceptual model. The model shows how these pressures change different aspects of the flow regime and plant habitats, and we describe how these changes affect the extent of the riparian zone and dispersal, germination, growth and competition of plants. Main consequences of the pressures we studied are the decrease of the extent of the riparian zone and a poorer state of the area that remains. This already results in a loss of riparian plant species and riparian functionality, and thus also threatens aquatic systems and the organisms that depend on them. We also found that the impact of a pressure does not linearly reflect its degree of ubiquity and the scale on which it operates. Hydropower and the climate crisis stand out as major threats to boreal riparian zones and will continue to be so if no appropriate measures are taken. Other pressures, such as forestry and different types of land uses, can have severe effects but have more local and regional consequences. Many pressures, such as non-native species and the climate crisis, interact with each other and can limit or, more often, amplify each other’s effects. However, we found that there are very few studies that describe the effects of simultaneously occurring and, thus, potentially interacting pressures. While our model shows where they may interact, the extent of the interactions thus remains largely unknown.

Introduction

Riparian zones are as important for ecosystems and their functioning, as they are vulnerable to a multitude of direct and indirect stressors caused by human activities ( Naiman et al., 2005 ; Stella and Bendix, 2019 ). Forming on the interface of land and freshwater, the riparian zone ( Figure 1 ) is often defined as the area between the low- and high-water mark along streams and lakes plus the part of the landscape that is above the high-water mark but still in direct exchange with the water table ( Naiman et al., 2005 ). Variation in regional and local geomorphology, and the consequent impacts of flow and sedimentation regimes, cause high heterogeneity in riparian zones through changing the recruitment of and dynamics within the riparian zone ( Vesipa et al., 2017 ). This heterogeneity is the basis for a species-rich plant community that contributes to local and regional diversity ( Nilsson and Svedmark, 2002 ; Sabo et al., 2005 ) and provides many other organisms with resources (e.g., Bennett et al., 2014 ; Johnson and Almlöf, 2016 ). Additionally, riparian zones and their vegetation fulfil a disproportionately large role in the functioning of fluvial landscapes, for example by physical and chemical buffering and cycling.

Figure 1. Vertical zonation in boreal riparian zones. Image by Lovisa Lind.

This review focusses on riparian vegetation in the boreal zone, an area that is characterised by a short growing season (3–6 months), cool summers, long, cold and snow-rich winters ( Pfadenhauer and Klötzli, 2020 ), and year-round precipitation with a peak in summer ( Beck et al., 2018 ). The boreal zone is delimited by the 10°C-isotherm in the north and the 18°C-isotherm in the south and stretches across North America, Scandinavia and Russia ( Pfadenhauer and Klötzli, 2020 ). The geology of the boreal zone with respect to the lithosphere is very variable and consists largely of continental crusts or sedimentary rocks from the Archean to the Cenozoic era ( OneGeology, 2020 ). Large parts of the boreal zone consist of mountains or are covered by mires, and most soils were shaped during the latest ice age and developed during the Holocene, on unconsolidated rocks such as loess, moraines and fluvio-glacial sediments ( Pfadenhauer and Klötzli, 2020 ). In boreal systems, year-round precipitation and relatively low temperatures assure year-round flow in most rivers, but variation between years can be large ( Woo et al., 2008 ). The melting of ice and snow at the end of winter marks the beginning of the spring flood, which is the peak flow in this climatic zone ( Lindström, 1993 ). Flow is lowest during summer, when evapotranspiration is highest, increases due to increased precipitation in autumn, only to decrease again when winter commences and rivers (partly) freeze. Freezing can take place on the water surface or from below, and the formation of ice dams can cause high water levels in the stream and flooding or ice formation in riparian zones ( Lind et al., 2014a ). Boreal riparian zones are thus characterised by a natural disturbance regime, in which parts of the riparian vegetation get scoured away by ice or flooding, and sediments are deposited annually ( Nilsson and Svedmark, 2002 ; Yarnell et al., 2015 ). These processes lead to a zonation of the vegetation ( Figure 1 ), with vegetation belts forming mainly based on the frequency and intensity of flooding. In addition, local factors, such as high connectivity with the groundwater table, can alter chemical cycling and circumstances in the riparian zone to the extent that plant species richness increases ( Kuglerová et al., 2014b ). Water and material are also received from upland habitat via surface runoff or subsurface flow, which means that the riparian zone is tightly connected with all of its surroundings and a hotspot for the conversion, transportation and storage of water and material in fluvial systems ( Naiman and Décamps, 1997 ).

The connectedness of riparian zones to its adjacent streams and upland habitats makes them, and their vegetation, vulnerable to changes in their environment. Riparian zones themselves have since long been exploited by people, for example for agriculture or herding, as a means to access water or for extraction of raw materials ( Naiman et al., 2005 ; Langston, 2013 ). However, increasingly intensive and large-scale activities, both in-stream and upland, such as damming and forestry, respectively, affect local and regional processes ( Bejarano et al., 2020b ; Kuglerová et al., 2021 ). These separate pressures, combined with global pressures such as the climate crisis ( Nilsson et al., 2013 ), have had, and continue to have, profound effects on the composition and functioning of riparian vegetation.

Scientific interest for riparian vegetation has increased over the past decades, to the point where it has become a rather well-described topic ( Dufour et al., 2019 ; Rood et al., 2020 ). Dufour et al. (2019) recognise, however, that the topic remains quite scattered, and found that the geographical and climatological spread of studies is far from equal: most work describes systems in North America and Europe, and in temperate or drier zones. Some areas are relatively understudied, such as parts of Russia and North America, because they are remote, (relatively) pristine or both, whereas other areas, such as boreal Sweden and Finland, receive relatively much attention. Studies often describe limited spatial and temporal scales and multiple stressors and pressures on riparian vegetation, a trend that is also recognised by other authors ( Stella and Bendix, 2019 ). That entanglement means that there is a lack of insight in what the individual effects of these pressures are, and which patterns are observed because of additive, synergistic or antagonistic effects ( Stella and Bendix, 2019 ).

Here, we aim: (1) to conduct and present a literature review ( Grant and Booth, 2009 ) on the state of knowledge on the different pressures on boreal riparian vegetation, and how these interact. We present a conceptual model ( Miro, 2021 ) in which we (2) visualise the effects of these pressures on the main processes that build up, maintain and break down riparian vegetation. We also (3) review the potential consequences of these changed processes on specific groups of plants. We used Google Scholar as a starting point to search recent and current, English-written scientific literature on pressures on boreal riparian vegetation and searched for additional literature by using the reference and citation lists of the initial results. This resulted in 145 original papers, 23 reviews, 12 reports and 2 books that describe these pressures and the ecological processes they affect or are expected to affect, or their effects on the vegetation itself. In total, we found 182 sources describing lab, field and modelling research on the effects of flow regulation for hydropower generation, flow regulation through channelisation, the climate crisis, forestry, land use change and non-native species on riparian vegetation and vegetation-shaping processes ( Stella and Bendix, 2019 ; Rood et al., 2020 ; Laudon et al., 2021 ; Singh et al., 2021 ).

Pressures and Their Effects on Boreal Riparian Vegetation

Flow regulation for hydropower generation.

Regulation of natural flow regimes for freshwater storage, flood control and transport is a widespread phenomenon ( Naiman et al., 2005 ). During the 20 th century, generation of energy has been the predominant reason for the building of in-stream infrastructure across the boreal zone. For example, flowing water is the third largest source for energy production in Sweden ( Energimyndigheten, 2020 ) and the largest source for energy production in Canada ( Natural Resources Canada, 2020 ). Nilsson et al. (2005) report that large river systems in boreal areas are relatively little affected by damming activities, when compared to those in other climatic zones, except in Sweden. Hydropower will also become more important in the future, as plans for an energy transition continue to be developed (e.g., Couto and Olden, 2018 ; Smokorowski, 2021 ). While hydropower is a renewable form of energy, its negative effects on natural ecosystems and on other ecosystem services, are many and well-documented (e.g., Renöfält et al., 2010 ; Tonkin et al., 2018 ). Hydropower infrastructure can be used for decades and causes altered geomorphology ( Englund and Malmqvist, 1996 ), changed flow regimes ( Nilsson and Svedmark, 2002 ) and decreased longitudinal connectivity ( Jansson et al., 2000b ) by dividing rivers in dam-to-dam ecosystem fragments rather than well-connected entities ( Wohl, 2017 ). Dams in run-of-the-river systems are often built at rapids to use the naturally present streambed height difference, and the rapids are often bypassed, causing them to fall dry ( Renöfält et al., 2010 ). The placement of dams at these rapids also means that upstream, lotic river fragments become lentic and experience little variation in water levels ( Figure 2 ). They turn into reservoirs, or reservoir-like river stretches, and while the water level fluctuations previously were large on a seasonal scale and small on a daily basis, there will be hardly any seasonal variation left after the regulation ( Englund and Malmqvist, 1996 ; Arheimer et al., 2017 ). There is a spectrum of flow regulation types and intensities with which hydropower is generated, with, for example, run-of-river being less intensive than hydropeaking ( Bejarano et al., 2018a ). Depending on the type of regulation, variation in flow is then higher on a weekly, daily or sub-daily basis than on a seasonal basis.

Figure 2. (Left) : View over the regulated Västerdal River, Sweden. Photo by LL. (Right) : A riparian zone undergoing hydropeaking in Ume River, Sweden. Photo by JH.

Hydropeaking ( Figure 2 ) is an often-applied form of water level regulation that leads to variation on a sub-daily basis. The frequent inundation caused by hydropeaking leads to soil waterlogging and submergence of plants that causes slow gas diffusion, rapid light attenuation and anoxia ( Armstrong et al., 1994 ; Bejarano et al., 2018a ). Next to that, hydropeaking leads to frequent water shortage, especially during low flow conditions, which also causes plant stress. These are quite general effects on plants, but hydropeaking can also have negative effects on plants’ germination success in the riparian zone, and their subsequent establishment, resulting in vegetation communities that are especially poor in flooding-intolerant plants ( Bejarano et al., 2020b ) and richer in competitive species ( Aguiar et al., 2018 ). Increased flow regulation leads to changes in germination, establishment and growth ( Bejarano et al., 2018a ; Greet et al., 2020 ), to a lower species richness locally ( Nilsson and Jansson, 1995 ) and lower abundance of all plant life forms in general, and of non-woody species in particular ( Bejarano et al., 2020a ). With a flooding regime that changes from high amplitude to high frequency, the riparian zone, whose width is determined by the spring flood, becomes narrower. Hence, another direct and local consequence of hydropower is the decrease in riparian space. While this is not reported in literature yet, this decrease may lead to a decrease of riparian species in a similar way as is predicted to happen with decreased riparian extent as a consequence of climatic change ( Ström et al., 2012 ; Jansson et al., 2019 ). These effects are visualised in Figure 3 , where “Flow regulation for hydropower regulation” affects the amplitude, timing, frequency and duration of floods, which in turn affects the extent of the riparian zone and processes therein, such as plant growth.

Figure 3. Conceptual model of the main potential effects of anthropogenic pressures on processes and habitats in the boreal riparian zone. Circles are pressures and pressure components, rectangles are intermediates, partially grouped as belonging to the flow regime (upper shaded rectangle) or the microhabitat (lower shaded rectangle), and squares are characteristics of and processes in the riparian zone. Lines and arrows represent effects and can be positive or negative. The effects of “Ice dynamics” change over time with initial intensification of the dynamics and eventual decrease with increasing climatic change.

These local changes in the riparian vegetation community are often (but not always, see Nilsson and Jansson, 1995 , and Jansson et al., 2000a ) reflected on the regional scale, with many studies demonstrating changes in riparian vegetation composition along regulated rivers, compared to free-flowing rivers (e.g., Nilsson et al., 1991 , 1997 ; Jansson et al., 2000b ). These changes, often in the form of decreased species richness within, but not necessarily across, river fragments, can be caused by dams physically disconnecting river fragments from each other (see “Infrastructure” in Figure 3 ). Each dam limits the transport of sediments and nutrients, which leads to upstream accumulation and a relative deficit downstream ( Hauer et al., 2018 ). The lower flow velocity downstream of the dam also decreases erosion and transport distance of materials, leading to a relatively larger deficit locally although local circumstances can change such patterns, such as abundant vegetation that retains sediments ( Vesipa et al., 2017 ). Dam infrastructure thus changes resource availability through multiple mechanisms, but it also affects biota and biotic populations through limitation of migration of organisms and dispersal of plant propagules ( Andersson et al., 2000b ; Mallik and Richardson, 2009 ; Renöfält et al., 2010 ). While many studies suggest that dams reduce species richness in and similarity between impoundments ( Andersson et al., 2000a ; Merritt and Wohl, 2006 ; Nilsson et al., 2010 ), there is evidence that hydrochory from within-impoundment sources can compensate for decreased dispersal between impoundments, leading to similar amounts of water-dispersed propagules in impounded and free-flowing rivers ( Jansson et al., 2005 ).

Timing and duration of flooding can also change vegetation composition and functional diversity ( Lozanovska et al., 2020 ). During a rise in discharge, vegetative propagules and seeds could become washed out and transported downstream, and later strand during decreasing discharge ( Bejarano et al., 2018a ). Sarneel et al. (2016) found that the timing of arrival of seeds does affect their chances for germination and growth, meaning that priority effects, i.e., the advantage that already present or early-arriving species have over later-arriving species, can put hydrochorous species at a disadvantage when compared to species that rely on other forms of dispersal. Hydrochory is extensively studied as a factor that affects local and regional species richness and hydrochorous species specifically ( Jansson et al., 2000b ; Merritt et al., 2010 ), and is one of the mechanisms driving riparian vegetation to become less species-rich and, probably, also functionally poorer with increasing regulation for hydropower purposes ( Poff and Zimmerman, 2010 ; Bejarano et al., 2018b ). While much of the work on riparian vegetation in boreal ecosystems has been done in northern Sweden, it is suggested that similar mechanisms drive changes in boreal riparian communities elsewhere ( Dynesius et al., 2004 ).

Given that regulation for hydropower generation, in whichever form, changes the processes that underlie the structure of riparian vegetation, its effects are diverse but tremendous. While it is difficult to generalise, there is sufficient literature (e.g., Nilsson et al., 1997 ; Jansson et al., 2000a ) to support the hypothesis that species and groups of species with specific traits decrease in abundance or become locally or regionally extinct. The fact that regulation is widespread and that it has both local and regional effects wherever it occurs, also means that its interactions with other pressures are manifold. These interactions will be discussed at the respective pressures throughout this paper.

Flow Regulation Through Channelisation

While some rivers are regulated for hydropower purposes, others have been adjusted for timber floating. From the mid-nineteenth century until the 1970s (parts of) many boreal rivers in Scandinavia were channelised to facilitate timber floating ( Gardeström et al., 2013 ). Boulders were removed and riverbeds were narrowed and smoothened ( Figure 4 ) to speed up downstream transport of logs ( Muotka and Syrjänen, 2007 ). This simplification of the channel morphology led to higher flow velocity and different flooding dynamics in the riparian zone ( Nilsson et al., 2005 ), and resulted in increased longitudinal connectivity whereas vertical and lateral connectivity decreased ( Wohl, 2017 ).

Figure 4. Bjurbäcken, a channelised tributary to Vindel River, Sweden. Photo by Jacqueline Hoppenreijs.

The fluvial effects of channelisation are often opposite to those of damming, as flow velocity increases and residence time of water in the channel thus decreases. The increased longitudinal connectivity also increases the distance over which sediment and, potentially, other material such as plant propagules, are transported in a certain amount of time ( Figure 3 ; Nilsson et al., 2010 ). We did not find any research that describes direct effects of increased longitudinal connectivity on the composition of riparian vegetation.

The degree to which the riparian zone is affected by in-stream changes also depends on the lateral connectivity between stream and riparian zone. The placement of boulders on riverbanks means that there is less space for riparian plants to establish in channelised streams ( Figure 3 ). Helfield et al. (2007) found that channelised streams have a lower frequency of low-intensity floods when compared to restored channels, meaning that these boulders have also hampered the deposition of sediment, nutrients and, probably, plant propagules. High flow velocity is usually related to decreased deposition (e.g., Wohl and Beckman, 2014 ), which means that riparian zones along channelised streams in boreal areas may be prone to relatively large depletion effects. This further decreases opportunities for establishment and growth of riparian vegetation. Decreased nutrient availability may affect some species more than others, which may cause shifts in riparian community composition. Erosion of the riparian zone is unlikely in parts of the channel where artificial channel structures are in place, but more likely in places without such structures, because of the higher flow velocity ( Kuglerová et al., 2017 ). The lower disturbance rate caused by the boulders on riverbanks likely also resulted in fewer open patches and thus even fewer opportunities for plant establishment ( Kuglerová et al., 2017 ). Therefore, there will be fewer plant propagules released from the riparian zone into the stream. Since spreading of plant propagules via water is an important driver of riparian plant species richness ( Jansson et al., 2005 ), it is likely that this has caused riparian vegetation to change during the years during which streams were channelised.

We do not know of any literature that looks at the effects of channelisation on pristine riparian vegetation. Knowledge on effects of channelisation on boreal streams is largely obtained through comparison of channelised with natural or restored streams. This lack of knowledge has resulted in few BACI (before-after-control-impact) designs being published (but see Nilsson et al., 2015a for some before-after studies on other taxa), and space-for-time substitutions (e.g., Hasselquist et al., 2015 ; Dietrich et al., 2016 ) are being used to overcome this lacuna. Most literature supports the hypothesis that restoration has or will have positive effects on riparian vegetation and, thus, that channelisation has had negative effects (e.g., Helfield et al., 2007 ; Hasselquist et al., 2015 ). Not all post-restoration data point consistently into the direction of complete recovery toward a state that resembles natural riparian vegetation, but some research suggests that aspects of natural streams were enhanced. For example, Kuglerová et al. (2017) found higher substrate availability along enhanced-restored reaches than along channelised streams. Depending on scale, restored streams had more or equally species-rich and species-even vegetation than channelised streams ( Helfield et al., 2007 ; Kuglerová et al., 2017 ), although other studies point out that riparian communities did not necessarily consist of more typical riparian species and that vegetation may need more time to recover than currently has been studied ( Helfield et al., 2012 ; Hasselquist et al., 2015 ; Nilsson et al., 2017 ).

Channelisation of boreal streams is a consequence of human activity that mainly, although not exclusively, happened in the past. Riparian vegetation is affected in stream segments that are still channelised, and seems to be recovering in streams that have been restored. Consistent, long-term monitoring is needed to increase understanding of the recovery process but, given the long recovery time, the current lack of understanding of this process should not be a reason to postpone restoration measures.

The Climate Crisis

The climate crisis, a self-inducing process caused mainly by burning of fossil fuels and deforestation, entails increased global warming, rising seawater levels and increasingly frequently occurring extreme weather events ( IPCC, 2014 ). These changes will lead to shifts in climatic zones within the next century ( Beck et al., 2018 ). Temperatures in northern Scandinavia, an area now classified as boreal in the Köppen-Geiger classification, will increase and result in less snowfall but more rain, and net higher precipitation ( Hoegh-Guldberg et al., 2018 ). The shorter freezing periods will lead to an earlier spring flood with reduced amplitude and duration, and higher discharge during autumn and winter ( Andréasson et al., 2004 ; Woo et al., 2008 ), effects that are not unlike those of hydropower regulation ( Arheimer et al., 2017 ). Interactions between the changed timing of floods, rising seawater levels and fluctuations therein may further change the flooding regime ( Kasvi et al., 2019 ). In general, discharge is projected to increase on the catchment scale ( Palmer et al., 2008 ). Whilst all these changes concern the entire boreal region, the magnitude of their effects on the hydrological regime differs on the regional and local scale ( Andréasson et al., 2004 ; Teutschbein et al., 2015 ).

The reduced seasonal variation in discharge will cause a narrowing of the riparian zone, which leaves less space for riparian species and enables the vegetation zones above and below the riparian vegetation to encroach river banks ( Figure 3 ; Nilsson et al., 2013 ; Jansson et al., 2019 ). Lower spring floods cause upland vegetation to expand and take over the upper parts of the riparian zone while higher autumn and winter flows enable aquatic and amphibious vegetation to do the same at the lower parts ( Ström et al., 2011 ). Ström et al. (2012) describe that pattern in more detail, predicting that riparian forest and shrubs decrease in area, that graminoid vegetation shifts upward and that amphibious vegetation expands. There will be large differences in the severity of these effects between riparian species ( Jansson et al., 2019 ). Some species will be affected but their occurrence in other than riparian habitat types may prevent population collapses and local extinction in the nearest future. In contrast, exclusively riparian species that are sensitive to changes in discharge are in more immediate danger, hence the predicted decrease of species richness of riparian vegetation ( Nilsson et al., 2013 ).

All species in the riparian zone are subject to the other immediate consequences of a changed climate, such as altered snow cover duration, groundwater tables or evapotranspiration rates ( Figure 3 ; Nilsson et al., 2013 ). Since these changes do not occur evenly throughout the year and will lead to species-specific responses, they will change riparian vegetation composition ( Sarneel et al., 2019b ). Shifting species-specific life histories and shifting biotic and abiotic regimes can occur in the same direction and at similar paces, but also in different directions and with different speeds. Non-matching shifts can lead to mismatches, or asynchrony, such as those reported for germination of riparian species and changed hydrological regimes ( Greet et al., 2011 ). Thus, these mismatches may interfere with key processes of vegetation development, such as dispersal, germination, growth and survival, and reproduction. Abiotic shifts may for example concern resource availability ( Perry et al., 2020 ) or hydrological processes ( Stella et al., 2006 ; Balke and Nilsson, 2019 ). Because different species are affected in different ways, riparian vegetation composition is likely to change in the future. Although changes can be predicted for some species (e.g., Balke and Nilsson, 2019 ; Perry et al., 2020 ), the effects of extreme events, interactions between species, large local variation and interactions with other pressures make it difficult to predict what future riparian zones will look like, but decreased taxonomic and functional diversity are to be expected ( Nilsson et al., 2013 ; Baattrup-Pedersen et al., 2018 ).

A process that may counteract this diversity decline in the near future is the expected increase in ice dynamics ( Figure 3 ). Despite the shorter winter season, northern boreal streams will go through the cycle of freezing and thawing more often because of higher winter temperatures ( Andréasson et al., 2004 ; Lind and Nilsson, 2015 ). High ratios of frazil and anchor ice formation are related to species-rich riparian vegetation ( Engström et al., 2011 ). The winter flooding and scouring of ice in the riparian zone create new patches for succession to start over ( Lind et al., 2014a ). In the long term, however, global warming will decrease ice formation to such an extent that ice dynamics will decrease, as will their positive effect on native species diversity in riparian zones ( Lind et al., 2014b ; Lind and Nilsson, 2015 ).

Interactions with other pressures can be a double-edged sword. While streams that are already deteriorated, for example because of hydropower, may be more sensitive to the negative effects of climatic change ( Palmer et al., 2008 ), the consequences of the climate crisis can lead to increased potential for hydropower ( Graham et al., 2007 ; Renöfält et al., 2010 ). There may also be possibilities for usage of hydropower infrastructure to mitigate the effects of climate change ( Arheimer et al., 2017 ) by mimicking natural flow regimes. Other authors conclude that the ways in which hydropower infrastructure and climate change may interact are potentially dangerous to their surroundings and can have adverse ecological effects, especially in a context of extreme events ( Palmer et al., 2008 ; Lejon et al., 2009 ). We illustrate the potential effects of simultaneously occurring pressures such as flow regulation for hydropower purposes and climatic change on one specific example, the riparian soil water table, in Box 1 . Hydropower infrastructure is known to limit dispersal and migration, which enables climate-related shifts in species distributions, that usually take place toward colder, upstream regions ( Nilsson et al., 2005 ). To our knowledge, there is no research that describes the effect that such infrastructure or channelisation has on the upstream dispersal of riparian plants, but Fink and Scheidegger (2021) show that connectivity along rivers is vital for riparian species to be able to disperse to future suitable habitats.

The climatic changes expected to occur in the boreal region ( Beck et al., 2018 ) imply that the environmental filter ( sensu Catford and Jansson, 2014 ), which currently hampers the establishment of non-native species, may shift. More specifically, higher temperatures may facilitate populations of already existing non-native species to expand or new populations to establish, and the corridor function of riparian zones can eventually cause further spread into upland habitats ( Nilsson et al., 2013 ). The same authors do not expect this to lead to an immediate loss of native species, although increased competition can lead to changes in species composition. While the riparian zone as a whole is vulnerable to invasions ( Rose and Hermanutz, 2004 ), the middle parts of reaches seem most susceptible ( Renöfält et al., 2005 ) and may thus show the effects of changed climatic filtering first. Another process affecting the potential success of non-native species is the expected increase in ice dynamics ( Lind et al., 2014b ) that will initially take place. Increasing ice dynamics may offer more opportunity for non-native species to establish in the area, potentially leading to displacement of native species. This eventual effect will decrease with time, as the increasing temperatures will stop the water from freezing all together.

Neither the expected changes in the separate and interacting pressures over time nor their combined effects are sufficiently quantified to make reliable predictions on how riparian vegetation will change on the local, regional or global scale. Although several studies cited here (e.g., Lind and Nilsson, 2015 ; Baattrup-Pedersen et al., 2018 ; Sarneel et al., 2019b ) cover multiple years, short-term changes found in such work cannot always be extrapolated to the medium or long term (see for example Blume-Werry et al., 2016 ). Much recent and valuable research aims to predict climatic changes (e.g., Arheimer et al., 2017 ) and the effects these may have on riparian vegetation, using qualitative ( Catford et al., 2013 ) or quantitative models ( Ström et al., 2012 ; Jansson et al., 2019 ). Much of this work, however, focusses on the general trends of increasing temperature and changing precipitation regimes. Extreme events, such as extreme floods, droughts or wind, are often overlooked ( Figure 3 ; Walsh et al., 2020 ) but will occur more frequently with increasing climatic change (e.g., Beniston et al., 2007 ) and are thought to impact already disturbed catchments more than non-regulated catchments ( Palmer et al., 2008 ). According to Van Oorschot et al. (2018) , extreme events that result in higher discharge, can lead to a shift of riparian vegetation toward upland in temperate ecosystems. On the other hand, their “drying scenario” suggests a shift of the riparian belt toward the stream. Bjerke et al. (2017) show that extreme events can have profound effects on boreal vegetation in general, and Nilsson et al. (2015b) predict direct and indirect negative effects on riparian vegetation.

Despite the uncertainties regarding the exact magnitude of climatic change, the pace with which it takes place and what role extreme events will come to play, the literature (e.g., Nilsson et al., 2013 ; Lind et al., 2014b ) points out that these changes will be reflected in riparian vegetation. Theoretical models and empirical evidence consistently point out that riparian species composition, and thus the functioning of riparian ecosystems, will change (e.g., Ström et al., 2012 ; Baattrup-Pedersen et al., 2018 ). Much remains unknown about how different pressures interact with each other, especially when it comes to pressures that operate on different scales. Future research and management should focus on combining knowledge on global, regional and local processes to tailor measures for specific areas.

Large parts of the boreal zone are covered with coniferous forest, which is one of the densest forest types in the world ( Crowther et al., 2015 ). This makes the parts that are within reasonable distance from society interesting from an economical point of view. Indeed, forestry has become the dominant human land use in the Scandinavian (see for example Östlund et al. (1997) , who describe the Swedish case) and North American ( Wells et al., 2020 ) parts of the boreal zone, and it is an important industry in Russia as well ( Leskinen et al., 2020 ). Forestry has replaced fire as the dominant disturbance regime in the upland habitat. Riparian vegetation, little affected by fires because of its proximity to water, may harbour trees that are relatively older than in the upland habitat, which makes them of higher interest for forestry ( Timoney et al., 1997 ). Forestry has thus become an important activity in both upland habitat and in the riparian zones, but forestry practices and management of streams within forestry areas are different in different countries, have changed over time, and will continue to do so (see for example Lazdinis and Angelstam, 2005 ). At the same time, forestry consists of many different phases that each have different effects on riparian zones ( Kuglerová et al., 2021 ). Its consequences are thus quite variable in time and space, and we will focus on the local effects of logging in the riparian zone and in upland habitat.

Logging in riparian zones leads to direct and indirect changes in vegetation composition, that occur both immediately and on longer time scales. With regard to the former, direct and immediate changes occur because of the removal of trees, which have been growing there for decades or even centuries ( Figure 3 ). Target species such as Larix spp., Picea spp. and Pinus spp. will make up a smaller part of the plant community immediately, as adult trees will be harvested ( Timoney et al., 1997 ). Logging also affects the understory vegetation, for example as a consequence of a changed microclimate following increased wind and sun exposure ( Figure 3 ; Chen et al., 1995 ; Berrigan et al., 2021 ). In addition to that, the planting of species that are interesting from an industrial perspective may lead to non-native tree species becoming invasive ( Richardson and Rejmánek, 2004 ; Marinich and Powell, 2017 ). These can outcompete native species or change the microhabitat. Other potential drivers of change are decreased evapotranspiration, leading to a higher water table and higher surface temperatures and, thus, altered nutrient cycling ( Foley et al., 2003 ; Luke et al., 2007 ). MacDonald et al. (2014) found that this does not necessarily lead to major shifts in riparian understory composition, although the limited duration of their study makes it difficult to predict medium- and long-term effects. They did find that species turnover can be high and that the perceived resistance to change probably relies on high nutrient availability and the natural flow regime. Such changed circumstances can favour certain species, such as ruderal, shade-intolerant and generalist species along headwater streams ( Newaz et al., 2019 ), over others. These effects may be stronger along streams of which both sides are logged and along larger streams, where there are no other trees to provide shading.

While riparian buffers can come in many forms, which also brings different functionality ( Kuglerová et al., 2020 ; Sonesson et al., 2020 ), all types of buffering imply some sort of limit to harvesting in the riparian zone. Limited logging and insufficient buffer width can still lead to changes in microclimatic conditions ( Jyväsjärvi et al., 2020 ; Berrigan et al., 2021 ), which means that vegetation can also change in buffers. Indeed, Oldén et al. (2019) found changes in riparian vegetation composition even in buffers in which no logging took place, although they did not find significant shifts for individual species. Contrastingly, Mallik et al. (2013) found no shifts in vegetation composition, but detected morphological changes in understory vegetation. These changes likely have to do with edge effects and emphasise the importance of location-tailored buffering rather than fixed widths ( Kuglerová et al., 2014a ). Poorly designed buffers also lead to rates of wind throw that are higher than natural wind throw rates ( Mäenpää et al., 2020 ) and can cause even more bank erosion than in clear-cuts ( Hylander et al., 2005 ), probably leading to increased degradation of riparian vegetation where the intention was to preserve it.

Changes in riparian vegetation composition following logging can occur faster or be more pronounced in combination with other pressures. For example, Newaz et al. (2019) found that forestry activities themselves led to increases in ruderal species. Open canopies, one of the consequences of logging, have been related to the presence of non-native plant species elsewhere, meaning that forestry may facilitate spread of non-native species in riparian zones ( Warren et al., 2015 ). Climatic change increases the possibility of non-native species invading an area as well and may lead to more non-native species becoming invasive in the riparian zone ( Rose and Hermanutz, 2004 ). Another consequence of a changed climate, combined with the homogeneous character of forests that are managed for industrial purposes, is that these forests become more susceptible to pests ( Folke et al., 2004 ) and fires ( Stine et al., 2014 ; Hessburg et al., 2019 ) which can also affect riparian vegetation. The climate crisis is expected to cause earlier starts of the fire season ( Stocks et al., 1998 ), higher fire intensity and larger fire areas ( Dale et al., 2001 ). Kilpeläinen et al. (2010) calculated an expected 20% increase in the annual frequency of forest fires in Finland by the end of this century alone. Even if pests or fires occur in upland habitat ( Figure 5 ) and not in the riparian zone itself, they will lead to organisms or species from the upland habitat seeking refuge in the riparian zone, thereby exerting pressure on its vegetation composition ( Tolkkinen et al., 2020 ).

Figure 5. Burnt forest patches after the 2018 forest fires on the right bank of Örån, west of Björkberg, Lycksele municipality, Sweden. ©Lantmäteriet.

Most forestry activities affect small streams that are not regulated for hydropower or timber floating purposes, although they may be channelised and ditched for drainage purposes ( Hasselquist et al., 2021 ). Whenever flow regulation and forestry do occur in the same area, the hypothesis that a natural flow regime supports the resistance of riparian vegetation against negative effects of forestry ( MacDonald et al., 2014 ) implies that riparian vegetation will undergo significant changes, as will its functioning. One function that may be altered by forestry, is the riparian zone’s nutrient and pollutant retention function. While forestry itself does not lead to extra production or deposition of pollutants, it can lead to increased mobilisation of elements such as mercury. These would otherwise remain retained in the riparian zone after having been emitted or deposited from other sources ( Bishop et al., 2009 ; Ledesma et al., 2018 ). Most literature describes increased mercury concentrations in the stream water ( Eklöf et al., 2012 ) and stream biota ( Lindqvist et al., 1991 ), and information on its effects on vegetation is scarce. While uptake of mercury via plant roots seems limited ( Lindqvist et al., 1991 ), some trace metals may be toxic to certain riparian plants ( Tolkkinen et al., 2020 ) and can limit their growth ( Påhlsson, 1989 ). Other chemicals, such as pentachlorophenol that was used to treat timber ( Naturvårdsverket, 2009 ), metals released from forest roads ( Kuglerová et al., 2021 ) or pesticides, can also have negative effects on vegetation ( Ranjan et al., 2021 ).

The effects of forestry on boreal riparian vegetation are understudied when compared to the scale on which forestry takes place. The focus of most research has been on the effects of logging on in-stream factors such as stream temperature and invertebrate communities via riparian zones, i.e., on their functions rather than their composition (e.g., Gundersen et al., 2010 ; Kuglerová et al., 2014a ). There seems to be relatively little research dedicated to the direct and indirect effects of logging practice on riparian vegetation, although some work has been done since this gap was recognised by MacDonald et al. (2014) . While changes in riparian microclimate seem consistent, the patterns in vegetation development are not, if found at all. This may have to do with the time-scale on which studies have taken place so far, which is merely a fraction of the forestry cycle, and with the variety of forestry practices, which is reflected in the research done on it.

Land Use Change

Land use changes, such as mining, agriculture, aquaculture and urbanisation, are major drivers of degradation of riparian ecosystems worldwide ( Naiman et al., 2005 ). In this section, we explore some of the threats that are less described in the literature, but potentially significant in the boreal zone. The geological composition of this area makes many of its terrestrial parts interesting for mining of minerals. While we will not go into detail in the different types of mines, it is worth mentioning that their normal activity and their proneness for accidental leakages make them a risk for the local and regional environment. The surroundings of mines have different water chemistry and more polluted sediments ( Leppänen et al., 2017 ), higher concentrations of pollutants in riparian soil ( Saint-Laurent et al., 2010 ) and accumulation of pollutants in the riparian food web ( Gerson et al., 2020 ) in comparison with sites distant from mines. Riparian soils, moss and vegetation around mines can contain high concentrations of pollutants, even after the mine has been terminated ( Qiu et al., 2005 , but see Bailey et al., 2002 ). Pollutants, for example in the form of heavy metals ( Påhlsson, 1989 ), can limit plant growth, and damage to vegetation in environments where multiple pollutants co-occur may be even larger than expected based on individually measured concentrations of pollutants ( Schipper et al., 2011 ). Most work in this field, however, is on the potential for phytoremediation rather than the specific effects of mining on boreal riparian vegetation composition. One study describes the abundance of plant species such as Salix spp., Carex nigra , C. rostrata , and C. vesicaria , and the overall species richness and composition in relation to metal concentrations and finds that species composition at the three tested sites related more to local topography than metal concentrations ( Husson et al., 2014 ). Next to the potential toxic effects of mining, its activities often cause changes in the topography and geomorphology of an area, thus reshaping surfaces or sedimentation patterns ( Figure 3 ; Naiman et al., 2005 ). It is therefore very unlikely that mining, through pollution or through changing ecological processes, would have no adverse effects on riparian vegetation on the local and the regional scale.

Agriculture and aquaculture are considered major drivers for plant extinctions ( Lughadha et al., 2020 ), and while riparian zones are among the habitats most likely subject to land use change, the authors specify that not all riparian zones are impacted equally. Although riparian zones are the most favourable areas for agriculture in boreal ecosystems, their relative remoteness and climatological limitations suggests that boreal riparian vegetation as a whole suffers relatively little from this type of land use on a regional scale ( Grizzetti et al., 2017 ). However, wherever agriculture takes place, removal of the vegetation through land cultivation or grazing leads to direct local destruction of the vegetation. Next to that, accompanying activities such as ditching ( Nybø et al., 2012 ; Jacks, 2019 ) and use of pesticides and fertilisers can cause long-term changes in hydrology ( Ledesma et al., 2018 ) and water chemistry ( Sponseller et al., 2014 ) and, thus, severely affect the local and regional riparian vegetation ( Figure 3 ; Lind et al., 2019 ). Some of these effects resemble the potential consequences of aquaculture ( Ahmed and Thompson, 2019 ) but, to our knowledge, this industry is far less widespread in boreal aquatic ecosystems. Under current circumstances, agriculture causes local, direct and indirect damage to boreal riparian zones and their vegetation, while regional, pollution-related damage is limited and global consequences pale in comparison to other pressures.

Urbanisation is a major pressure on riparian ecosystems all over the world ( Naiman et al., 2005 ; Gurnell et al., 2007 ). While the degree of urbanisation in the boreal countries has increased and is likely to continue to increase ( United Nations Department of Economic and Social Affairs Population Division, 2018 , but see Boverket, 2019 ), the boreal zone is one of the least populated land areas on Earth ( Center for International Earth Science Information Network [CIESIN], 2018 ) and the degree of urbanisation is thus relatively low ( Figure 6 ). We can distinguish between direct effects of urbanisation on riparian vegetation, through the clearing of riparian zones ( Walsh et al., 2005 ; Wheeler et al., 2005 ), and indirect effects, through processes that form the riparian zone ( Figure 3 ). For example, an increase of hard surface area leads to a decrease in soil permeability, which is likely to change the flow regime in the form of more extreme high flows ( Arheimer and Lindström, 2019 ), a mechanism that is probably similar in boreal ecosystems. Urbanisation is also known to lead to lower groundwater tables, which are the most likely cause for different riparian species composition ( Groffman et al., 2003 ), and run-off that is richer in nutrients than in non-urban areas ( Sponseller et al., 2014 ). Next to that, polluted run-off from urban spaces can be problematic for aquatic and riparian life, but seems limited in countries such as Sweden and Finland ( Grizzetti et al., 2017 ). Changed hydrology in urban riparian zones may lead to higher proportions of non-native species, at a cost of riparian plant diversity ( Burton et al., 2005 ). To our knowledge, however, there is no information on the expected effects of increasing urbanisation on riparian vegetation in boreal ecosystems, and the challenges in this specific field of research are manifold ( Nilsson et al., 2003 ). While large parts of the boreal zone are still not accessed or inaccessible ( Leskinen et al., 2020 ), the effects of urbanisation on riparian vegetation can be profound (e.g., Décamps et al., 1988 ; Aguiar and Ferreira, 2005 ) and may thus increase in the future.

Figure 6. Riparian land uses in the surroundings of Oulu, Finland. Based on the Copernicus Land Cover/Land Use classification ( Forslund, 2012 ) and the VHR2018 dataset from the European Environmental Agency. ©OpenStreetMap, and the GIS user community.

Increased human inhabitation of an area leads to more infrastructure and exurban activities, such as recreation by the area’s inhabitants and tourists. More and more intensively used infrastructure can lead to pollution in riparian zones, for example through the use of road salt, which can affect plant growth ( Stoler et al., 2018 ). Recreation can be non-consumptive and consumptive ( Naiman et al., 2005 ; Schafft et al., 2021 ), the former including activities such as camping and photography, and the latter fishing and firewood cutting ( Poff et al., 2011 ). All of these activities imply some kind of access which may lead to trampling of herbs and damage to seedlings or young shrubs and trees, which can advantage disturbance-adapted species ( Manning, 1979 ). Trampling, which has been studied in alpine and northern ecosystems, may have negative effects on vegetation cover ( Monz, 2002 ) and the abundance and richness of plants ( Jägerbrand and Alatalo, 2015 ). Although some communities will recover from light to moderate levels of trampling ( Monz, 2002 ), other communities may take decades to regenerate after severe trampling ( Willard et al., 2007 ). Next to that, diseases or non-native species can be introduced to an area, and consumptive activities may lead to decreases of specific species (groups) ( Poff et al., 2011 ). While none of these activities and their effects are described for boreal riparian ecosystems, the way they are for other areas (e.g., Madej et al., 1994 ), their damaging mechanisms are most likely the same ( Figure 3 ). Reachability of locations is a prerequisite for recreations to take place, and pressure by recreation is thus probably more common in Scandinavia than in Russia or North America. Moreover, recreational activities usually take place along larger stream orders ( Riis et al., 2020 ; Arif et al., 2021 ), which means that they do not affect headwaters and low-order streams that represent the largest part of the catchment and thus do relatively little damage to riparian vegetation on the regional and global scale.

Non-native Species

Riparian zones are thought to be sensitive to invasions by non-native species because of their direct connection to aquatic ecosystems (e.g., Nilsson et al., 2010 ), relatively intensive anthropogenic usage (e.g., Richardson et al., 2007 ) and because of their instable, disturbance-prone character (e.g., Naiman et al., 2005 ). The natural disturbance regime, consisting of processes such as ice-scouring or flooding events, make riparian zones relatively patchy and open for secondary succession and, thus, for the establishment of new individuals and new species. Not all non-native species become invasive, and factors such as the type of disturbance ( Jauni et al., 2015 , but see Ström et al., 2014 ), species characteristics ( Ni et al., 2021 ) and a system’s invasibility ( Lonsdale, 1999 ) codetermine the effects a species may have on an ecosystem. Here, we follow the definition of Mack et al. (2000) and define invasive non-native species as species that invade a new area in which they establish and form a threat to the abiotic or biotic environment.

Invasion and spread of non-native plant species occurs in riparian zones worldwide (e.g., Planty-Tabacchi et al., 1996 ; Rose and Hermanutz, 2004 ; Hejda and Pyšek, 2006 ; Ronzhina, 2020 ). While riparian vegetation in northern Scandinavia is usually described as primarily native ( Nilsson, 1999 as cited in Ström et al., 2011 ), this is in stark contrast with findings in other boreal regions (e.g., in North America) where Dynesius et al. (2004) reported up to 9% non-native species in riparian plant communities. Boreal riparian zones are usually low-competition ecosystems, which is why non-native plants may be able to outcompete native species, and lead to significant shifts in vegetation composition ( Figure 7 ). Common traits of invasive species, such as high biomass production, large stature, nitrogen fixation or high transpiration rates, may in turn change community structure, such as vertical biomass distribution, suppress processes such as germination, and affect ecosystem fluxes and functions, such as soil N or water availability ( Figure 3 ; Akamatsu et al., 2011 ; Simberloff, 2011 ; Ruwanza et al., 2013 ; Catford and Jansson, 2014 ). In addition, non-native invertebrates, fungi and micro-organisms can become pests, parasites or pathogens and damage native vegetation as such ( Kominoski et al., 2013 ; Lapin et al., 2021 ).

Figure 7. American Skunk-cabbage ( Lysichiton americanus ) along Klokkarhyttebäcken, Sweden. Photo by Owe Nilsson.

Non-native species do already occur in Scandinavian boreal riparian zones, but are not considered invasive ( Dynesius et al., 2004 ; Nilsson et al., 2013 ). Certain types of land and water use are considered a catalyst for the establishment of non-native species. Hydropower is the most important of such land uses in the boreal area, as it causes an increase of bare patches or creates new habitat by changing the flow regime and may thus enable non-native species to establish ( Tickner et al., 2001 ; Bejarano et al., 2020b ). Even species planted for forestry purposes can be considered invasive when they are planted in, or spread into, riparian zones ( Richardson and Rejmánek, 2004 ; Kominoski et al., 2013 ). The relatively limited character of other human activities, resulting in low propagule pressure ( Keller et al., 2011 ), may be a reason for the relatively minor problems with non-native species in boreal zones as compared with agricultural landscapes in Central Europe with the highest proportions of established species. Future changes, such as increased climatic change in combination with human activity, may however, increase the risk of further disruption of native riparian plant communities. Predictions on risks of establishment of terrestrial species show that large parts of boreal Scandinavia are under medium to high risk during the 21 st century, and boreal North America and Russia under low to high risk ( Early et al., 2016 ), but it is unclear in how far these projections can be used for aquatic species. Although projections of climatic change are not equal across the entire Scandinavian boreal zone, the generally higher winter temperatures and increased annual precipitation will support a different type of vegetation than has been found during the past centuries ( Nilsson et al., 2013 ). Other changes in the hydrological regime, such as a lower and earlier spring flood, will decrease the extent and change vegetation composition of the riparian zone ( Ström et al., 2012 ). While a lower spring flood may mean that lower numbers of non-native plant propagules are deposited on the riverbanks, we expect that the narrowing of the riparian zone can also lower the input of native plant propagules. The expected climatic changes will enable new species to disperse into the area, adventive species to establish actual populations and newly established populations to spread locally and fill distribution gaps. This can lead to non-native plant species becoming invasive in boreal riparian zones, and thus changing their composition and functioning.

Already degraded riparian vegetation is likely more vulnerable to the competitive capacities of non-native species than undisturbed riparian vegetation ( Rose and Hermanutz, 2004 ; Zelnik et al., 2015 ; Pattison et al., 2017 ). Established populations of invasive non-native species in the riparian zone may also present a risk for the surrounding landscape. While this may be an unsuitable matrix for species spread in itself, it may be prone to penetration by non-native species through increased propagule pressure from the riparian zone ( Richardson et al., 2007 ). Another, yet to be explored facet of riparian connectivity is the effect that restoration of hydropower-regulation may have on riparian vegetation. Restoration measures, such as the (partial) removal of dams, have the potential to increase the risk of spread of non-native plant species, especially around reservoirs ( Shafroth et al., 2002 ; Tullos et al., 2016 ). Studies from temperate regions show that these problems may be small ( Lisius et al., 2018 ; Ravot et al., 2020 ), but we do not know of examples from the boreal region.

Each of the anthropogenic pressures described in the sections above affects boreal riparian vegetation or parts of it through one or more biotic or abiotic ecosystem processes. We have visualised the pressures in a conceptual model, subdivided in components (smaller circles, such as “temperature,” “precipitation,” and “extreme events” as components of the climate crisis) if these can have different effects ( Figure 3 ).

The model indicates the effects of these pressures on different parts of the ecosystem and ecological processes that build the vegetation community and constitute a plant’s life cycle. The flow regime, the major driver behind riparian dynamics ( Lytle and Poff, 2004 ), is affected widely but not everywhere, through adaptations for hydropower ( Palmer et al., 2008 ), and globally as a consequence of the climate crisis ( Nilsson et al., 2013 ). The application of hydropower leads to direct destruction of riparian zones and undermines many of the processes that riparian vegetation depends on in the long term. On the other hand, anthropogenic climatic change is a process that is only just beginning, but expected to become more extreme in the future ( Arheimer et al., 2017 ). Forestry and other land uses can have profound direct effects locally and on the short term, for example when riparian vegetation is removed. They also have more regional, indirect and long-term effects such as pollution of water, increased insolation that damages shade-tolerant understory species, and wind throw in poorly designed buffer zones ( Lind et al., 2019 ; Kuglerová et al., 2021 ). In the past, forestry and channelisation for transport of timber, have led to more negative effects in riparian zones than other land uses. Even if other types of land uses become more widespread, it is likely that forestry practice in its current form will stay the main land use damaging riparian vegetation. Other land uses are, together with climatic change, the main drivers for the spread of non-native species into boreal riparian zones and may lead to a level of competition where native riparian plant species get outcompeted ( Dynesius et al., 2004 ).

While many of these changes can affect all species to some extent, e.g., when non-native species take up physical space at the cost of native species or when all riparian vegetation is removed for the placement of a dam, it is also possible to identify groups that are affected more than others. Such specific responses are described at the respective pressures that are found to cause them. Recognising where a pressure or a set of pressures can affect ecological processes or characteristics can open the door to more appropriate research questions and more adequate management. We illustrate the complex web of interactions between effects of multiple pressures on a single factor or process with one example, the riparian soil water table, in Figure 8 and Box 1 .

BOX 1. Flow regulation for hydropower limits the amount of water that riparian soils take up during spring by decreasing the spring flood ( Figure 8 ). Simultaneously, the climate crisis changes soil moisture through altered temperature patterns, which cause different evapotranspiration rates. In addition, forestry can lead to larger canopy gaps that cause more evapotranspiration, or forestry or agricultural ditches can increase discharge to the extent that water disappears from the riparian soil. These effects will change the riparian soil water table during at least some months of the year and at least in parts of the catchment, although changed precipitation patterns as another aspect of the climate crisis can counteract them to some extent.

Figure 8. An example of application of the conceptual model on the riparian soil water table. Pressures and processes that affect or are affected by the soil water table are marked in red.

The soil water table then codetermines the extent to which plants successfully germinate and grow, in a riparian zone of which it also codetermines the extent ( Figure 8 ). Not all species or species groups react to these changes in a similar manner, and in this example riparian or flooding-tolerant species that are drought-sensitive or demand shading, may experience reduced germination success and growth, and may eventually be outcompeted by ruderal or shading-intolerant species.

Our conceptual model brings together a wide range of anthropogenic pressures and their effects on riparian vegetation. With the knowledge currently available and the complex and variable way in which the pressures interact, quantifying their effects remains difficult for the foreseeable future. Simultaneously occurring pressures in an ecosystem, already difficult to quantify when considered separately, can have effects that counteract each other, work in an additive or even synergistic way or cause systems to cross thresholds relatively unexpectedly ( Stella and Bendix, 2019 ). Another difficulty is that most research that is available to the English-speaking community comes from Sweden and, to a lesser extent, from Finland and English-speaking Canada, and may not be directly applicable in Norway or Russia. The complexity of the effects of pressures makes timely research on relevant spatial scales more important than ever, especially in understudied areas.

In this review, we show that pressures occur simultaneously on local, regional and global scales. Their effects are not the same across these scales, and are not necessarily similar within scales, when measured at different locations or different times. This added variation is caused by interactions with the local and regional environment ( Bendix, 1994 ; Polvi et al., 2020 ). Legacies from past disturbances can also change how riparian vegetation develops ( Sarneel et al., 2019a ; Janssen et al., 2020 ), and makes that effects found in one catchment cannot always be extrapolated in space or time. This emphasises the importance of increasing the understanding of local and regional hydrogeomorphology and past disturbances in an area.

Understanding the history of an area can also help delimit an appropriate time scale for research and restoration plans. There are many practical, financial and, eventually, political factors that limit the possibilities of collecting data over a longer period of time ( Courchamp et al., 2015 ). That complicates the correct interpretation of some changes, for example, when comparing the effects of competition, which may be more pronounced on the short term, while other processes, such as dispersal, may only be measurable on a long-term scale. Inertia thus complicates ecological research, and puts another emphasis on the importance of long-term research efforts ( Turner and Gardner, 2015 ). And while rare examples of “ecosystem experiments” do occur ( Stella and Bendix, 2019 ), researchers do not always get the opportunity to conduct control measurements before a pressure-exerting activity starts. Intensifying collaborations between managers and researchers is as mutual a responsibility as it will be helpful to the advancement of the field. Studies such as those of Pickett (1989) and Nilsson et al. (2015a) acknowledge the difficulties with measuring control sites, or groups, in ecology and the need for more long-term research.

Another challenge is the identification of the right response variable. Both from a functional and from a conservation point of view, knowing which species will be affected can be more valuable than knowing that species will be affected. While our analysis includes species and groups that are very likely to react to anthropogenic pressures, it is difficult to be more specific. Not many studies have focussed on specific species or groups in the past, although there is a shift in recent years (see for example Baattrup-Pedersen et al., 2018 and Bejarano et al., 2018b ). In addition, species composition differs across the boreal zone and one specific species may react differently to pressures in different communities. We have therefore included in the text (1) species groups specifically mentioned in the literature and (2) groups with certain physiological, phenological or ecological characteristics that most likely are affected similarly across the boreal zone. Shifting the focus beyond species identity by analysing functional response or effect traits ( Truchy et al., 2015 ) can help increase the understanding of physical and ecological processes affecting the riparian zone and simultaneously increases the applicability of results across spatial and temporal scales.

We conclude that, while much progress has been made in the field of riparian research, the current levels of pressure call for a greater sense of urgency within the field and general governmental, scientific and societal practice. This review identifies forestry and hydropower as amply proven pressures on boreal riparian vegetation, and the literature shows that there are many ways in which involved actors can mitigate the negative effects of these industries, whereas the climate crisis calls for global action and involves a wider range of actors. Our conceptual model is also meant to function as a stepping stone for researchers and managers alike to explore mechanisms that could be relevant for their work. It will hopefully contribute to a better understanding of processes and interactions within boreal riparian ecosystems, and the way they are affected by external and internal pressures.

Author Contributions

JH, LL, and RE conceived the idea for the study and designed its structure. JH wrote the first draft. LL and RE made comments and additions to the manuscript. All authors read and approved the final manuscript.

The authors were funded by the Department of Environmental and Life Sciences at Karlstad University. Open access was funded by Karlstad University.

Conflict of Interest

The authors declare that the research was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest

Publisher’s Note

All claims expressed in this article are solely those of the authors and do not necessarily represent those of their affiliated organizations, or those of the publisher, the editors and the reviewers. Any product that may be evaluated in this article, or claim that may be made by its manufacturer, is not guaranteed or endorsed by the publisher.

Acknowledgments

The authors would like to thank their colleagues from the Department of Environmental and Life Sciences, Karlstad University, for valuable input and feedback on a rudimentary version of the model. The authors would also like to thank editor BK and the two reviewers for their comments on the manuscript.

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Keywords : riparian vegetation, boreal, hydropower, forestry, climate change, land use change (LUC), invasion, ecosystem interactions

Citation: Hoppenreijs JHT, Eckstein RL and Lind L (2022) Pressures on Boreal Riparian Vegetation: A Literature Review. Front. Ecol. Evol. 9:806130. doi: 10.3389/fevo.2021.806130

Received: 31 October 2021; Accepted: 22 December 2021; Published: 31 January 2022.

Reviewed by:

Copyright © 2022 Hoppenreijs, Eckstein and Lind. This is an open-access article distributed under the terms of the Creative Commons Attribution License (CC BY) . The use, distribution or reproduction in other forums is permitted, provided the original author(s) and the copyright owner(s) are credited and that the original publication in this journal is cited, in accordance with accepted academic practice. No use, distribution or reproduction is permitted which does not comply with these terms.

*Correspondence: Jacqueline H. T. Hoppenreijs, [email protected]

† ORCID: Jacqueline H. T. Hoppenreijs, orcid.org/0000-0002-4284-5453 ; R. Lutz Eckstein, orcid.org/0000-0002-6953-3855 ; Lovisa Lind, orcid.org/0000-0002-7212-8121

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Research Journal of Forestry

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Research Article

Diversity, relative abundance and distribution of avian fauna in and around wondo genet forest, south-central ethiopia.

Received: August 19, 2016;   Accepted: November 14, 2016;   Published: December 15, 2016

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Introduction, materials and methods.

SIGNIFICANCE STATEMENTS

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Species Composition , Diversity and Richness in Understanding Threats on Biodiversity Conservation of Philippine Native and Indigenous Species of Trees

The inventory and documentation of the tree species employed the transect method of ecological studies in determining species composition, diversity and richness of a forest sanctuary. The existing number of native and indigenous trees and the receding space for their growth due to the encroachment of fast spreading exotic plant species explain the threat to biodiversity of floral resources of the country. Consequently, the identification of priority species of native and indigenous trees for biodiversity conservation was given focus in the study because of their receding growth compared with invasive exotic trees. The study also points to the prevailing classification in the IUCN Red List of the country’s priority native and indigenous tree species as either CR-critically endangered or VU –vulnerable. The study site being a forest reserve and natural sanctuary plays important role in keeping these tree species protected and conserved for biodiversity and act as natural gene bank of...

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The Genus Artemisia : A 2012–2017 Literature Review on Chemical Composition, Antimicrobial, Insecticidal and Antioxidant Activities of Essential Oils

Essential oils of aromatic and medicinal plants generally have a diverse range of activities because they possess several active constituents that work through several modes of action. The genus Artemisia includes the largest genus of family Asteraceae has several medicinal uses in human and plant diseases aliments. Extensive investigations on essential oil composition, antimicrobial, insecticidal and antioxidant studies have been conducted for various species of this genus. In this review, we have compiled data of recent literature (2012–2017) on essential oil composition, antimicrobial, insecticidal and antioxidant activities of different species of the genus Artemisia . Regarding the antimicrobial and insecticidal properties we have only described here efficacy of essential oils against plant pathogens and insect pests. The literature revealed that 1, 8-cineole, beta-pinene, thujone, artemisia ketone, camphor, caryophyllene, camphene and germacrene D are the major components in most of the essential oils of this plant species. Oils from different species of genus Artemisia exhibited strong antimicrobial activity against plant pathogens and insecticidal activity against insect pests. However, only few species have been explored for antioxidant activity.

1. Introduction

Aromatic and medicinal plants are important sources of secondary metabolites, which have a wide range of applications in control of plant and human diseases, cosmetics, as well as in the pharmaceutical industry [ 1 ]. In the plant kingdom, family Asteraceae is endowed with essential oil-yielding plants, and among these plants, the genus Artemisia occupies top position for its bio-prospection. The genus consists of small herbs and shrubs, found in northern temperate regions and comprises of about 500 species from South Asia, North America and European countries [ 2 ]. Species of the genus are called by the common names mugwort, wormwood and sagebrush. Due to presence of terpenoids and sesquiterpene lactones, most of the species possess strong aromas and bitter tastes, which discourage herbivory, and may have had a selective advantage [ 3 ]. These species have wide and varied applications in plant and human disease control and in the pharmaceutical industry. There are several species of Artemisia that have been investigated as antimicrobial, antioxidant, cytotoxic, insecticidal, repellent and anticonvulsant agents [ 4 , 5 ]. Although a review on the genus Artemisia was published by Abad et al. [ 2 ] on the chemical composition, ethanopharmacological and biocidal activity of essential oils, they took the data from 2000–2011, and mainly focused on human pathogens. Recently, another review compiled by Al-Snafi [ 6 ], which focused only on A. campestris , revealed several pharmacological activities, such as antimicrobial, antioxidant, cytotoxic, insecticidal, antivenomous, and many other pharmacological effects. In this review, we have compiled the data from 2012–2017 on chemical composition, antimicrobial, insecticidal and antioxidant activities of Artemisia species. Regarding antimicrobial and insecticidal activities of essential oils, here, we have mainly focused on pathogens and pests of plants and, since Abad et al. [ 2 ] did not describe earlier literature on antimicrobial and insecticidal activities of Artemisia oils on same aspects and also antioxidant activity, so in this review we have also covered the literature from 2000 onwards on these aspects.

2. Chemical Composition of Essential oils of Artemisia Species

Plant essential oils are volatile in nature and consist of a complex mixture of monoterpenes and sesquiterpenes, which give strong odor to the essential oils. These essential oils are extracted from plants by various methods such as steam or hydro-distillation methods and are frequently being used in the natural product laboratory [ 7 ]. Essential oils are composed of more than 60 different components in different concentrations; among them few have higher amounts of composition. From time to time, the chemical composition of essential oils of the genus Artemisia has been studied by researchers from the different regions of the world. The essential oil composition of genus Artemisia investigated during 2012–2017 is reported in Table 1 . The table shows that investigator used leaf, stem, areal part and inflorescence for essential oil extraction and GC and GC/MS methods for the chemical analysis. 1, 8-Cineole, beta-pinene, thujone, artemisia ketone, camphor, caryophyllene, camphene and germacrene D were the major components reported in the essential oils of Artemisia species ( Table 1 ). The table also shows that the composition of essential oil of the same species varied in different investigations depending upon a change of geographical origin. Variation in the volatile components of these plants may occur during plant ontogeny or growth at different altitudes. However, few chemical constituents were restricted to limited species. For instance, methyl chavicol was only reported in higher amounts in A. dracunculus , piperitone in A. judaica, capillene in A. stricta and chamazulene in A. arborescens L, artedouglasia oxide in A. stelleriana . Most of the investigations into the chemical composition of essential oils were from Iran, followed by India and China.

A literature report from 2012–2017 on chemical composition of Artemisia oils from different geographical regions (Plant part: AP: aerial parts; F: flowers; FH: flower-heads; L: leaves; B: Buds)

3. Antimicrobial Properties of Artemisia Essential Oils

The interest in using essential oils as an antimicrobial agent is increasing mainly due to their natural origin, wide spectrum of activity and their GRAS (Generally Recognized as Safe) status. Since earlier reviews published on the Artemisia oils only described bioactivity against human pathogens, here we have illustrated literature from 2000 onwards on Artemisia oils only against plant pathogens. Essential oils from the Artemisia species have been explored for their antimicrobial properties against several bacterial and fungal plant pathogens by using different methods such as the disc diffusion method, agar dilution method, poison food method and inverted Petri plate method depending upon the nature of fungal and bacterial species [ 74 ]. In our laboratory bioassay, A. nilagirica oil performed potent results with postharvest pathogens of table grapes. Oil exhibited 100% mycelia inhibition against Aspergillus flavus , A. niger and A. ochraceus and 0.29 and 0.58 μL/mL fungistatic and fungicidal values, respectively, were noticed for all the fungal species. Oil (1.6 μL) completely suppressed the growth and mycotoxin (AFB1 and OTA) secretion of aflatoxigenic and ochratoxigenic strains of Aspergillus . Fumigation of oil (300 μL) was found to protect 1 kg of table grapes and enhanced the shelf life for up to 9 days. Thus, this oil can be used as grape protectant from fungal spoilage [ 75 ]. Sati et al. [ 18 ] examined in the laboratory that A. nilagirica oil is effective against root rot pathogens with ED 50 (effective dose) values against Rhizoctonia solani , Sclerotium rolfsii and Macrophomina phaseolina were 85.75, 87.63 and 93.23 mg/L, respectively. This oil also showed fungicidal activity against Colletotrichum fragariae , C. gloeosporioides , and C. acutatum and Artemisia caerulescens subsp. densiflora (Viv.) oil against Aspergillus , Alternaria and Fusarium species [ 26 , 76 ], whereas A. maritima oil has poor mycelial growth inhibition. In the three tested Artemisia ( A. scoparia , A. sieberi and A. aucheri) oils against soil-borne pathogens from Iran, A. aucheri and A. sieberi oils proved strong antifungal inhibitors with 41.406 μL/L EC 50 (effective concentration) for A. aucheri oil against Rhizoctonia solani , while A. sieberi oil showed 121.798 μL/L EC 50 with MIC value (minimum inhibitory concentration) 250 μL/L against R. solani . Artemisia sieberi oil was also fungistatic against Tiarosporella phaseolina (1000 μL/L), Fusarium moniliforme (750 μL/L) and F. solani (750 μL/L) with EC 50 values 203.419, 211.072 and 188.134 μL/L, respectively [ 77 ]. Essential oils of A. arborescens also reported as fungicidal agent against R. solani at 12.5 μL/20 mL [ 22 ]. Essential oil isolated from A. absinthium from different geographical regions showed significant antifungal activity (ED 50 0.5 μg/mL) against F. oxysporum and F. solani [ 78 ]. All these plant pathogens have a wide range of hosts including chick pea, mungbean, urdbean, soyabean etc., so their study suggests that these oils can be used as seed treatment for the control of these phytopathogenic fungi infecting agricultural crops. In the laboratory bioassay of Badawy and Abdelgaleil [ 79 ], A. monosperma oil was examined as a strong mycelial growth inhibitor of Alternaria alternata , Botrytis cinerea , F. oxysporum and F. solani and EC 50 values reported were 54, 111, 106 and 148 mg/L, respectively. However, they noticed that oils of A. judaica and A. monosperma caused highest spore germination inhibition of F. oxysporum at EC 50 values 69 and 62 mg/L, respectively . The growth of Sclerotinia sclerotiorum , a crown rot pathogen was inhibited by A. santonica , A. pontica , A. annua , A. austriaca , A. dracunculus , A. lerchiana , A. vulgaris and A. vulgaris var. pilosa oils at MIC of 2400 μL/L. However, A. abrotanum showed 1200 μL/L MIC value against S. sclerotiorum , while A. scoparia did not exhibit complete mycelial inhibition [ 80 ]. Some Artemisia oils showed weak antifungal activity against plant pathogens i.e., A. proceriformis oil showed a poor efficacy (MIC100 > 1.5 mg/mL) on Septoria glycine and other phytopathogens tested [ 57 ] which shows that this oil cannot be recommended as a plant protection product against said phytopathogens. This strong and poor effect of essential oil of different species of the same genus may be due to chemical composition and their synergistic effect. In another study, A. campestris oil showed a potent antifungal agent against F. graminearum , Penicillium citrinum , P. viridicatum and Aspergillus niger . The MICs reported were 1.25 μL/mL ( v / v ) for F. graminearum , while MFC for all fungal species exceeded 20 μL/mL [ 53 ]. The MIC of A. stricta oil was reported as 0.625 mg/mL against A. flavus followed by A. niger and Sporothrix schenckii [ 52 ]. Similarly, essential oils of A. absinthium showed 84 and 91 μg/mL MIC values for P. chrysogenum and A. fumigatus respectively [ 23 ].

In addition to essential oil screening, chemical compounds isolated from different species of Artemisia have also been evaluated against plant pathogens [ 81 ]. 5-phenyl-1,3-pentadiyne and capillarin isolated from A. dracunculus oil showed fungicidal activity against C. fragariae , C. gloeosporioides , and C. acutatum [ 82 ]. Dadasoglu et al. [ 74 ] assessed some chemical constituents like camphor, caryophyllene oxide, linalool, 1, 8-cineole, teminen-4-ol, borneol and α-terpineol isolated from A. absinthium , A. santonicum and A. spicigera oils against plant pathogenic bacteria and fungi. The MIC value of linalool was in the range of 50–110 mg/mL, terpinen-4-ol, 60–110 mg/mL for Xanthomonas campestris pv. vitians RK-Xcvi; α-terpineol-8, 60–70 mg/mL for Pseudomonas cichorii RK-166, P. huttiensis RK-260, P. syringae pv. syringae RK-204 and X. axonopodis pv. vesicatoria . Other chemical compounds such as caryophyllene oxide, bomeol, camphor and 1, 8-cineole did not show activity against any of the pathogens. Additionally, camphor, 1, 8-cineole and chamazulene isolated from A. absinthium oil from the Turkish population has been described as an effective fungicidal agent against wilt fungi F. solani and F. oxysporum [ 83 ] and from Uruguay the same species rich in thujone showed potent fungicidal activity against Alternaria sp. and Botrytis cinerea [ 84 ]. Thus, efficacy of Artemisia oils may be due to the presence of these chemical constituents and these chemical constituents can be used as potential antimicrobial agents against said pathogens [ 74 ]. Mycologists assumed that these chemical constituents present in the essential oils cause degeneration of fungal hyphae result in potent antifungal activity. Chemical compounds of essential oils dissolved in the membranes and therefore increase the permeability of the cell membrane, resulting in membrane swelling and reduction of membrane function [ 85 ]. Additionally, essential oils penetrate the cell walls of fungi due to their lipophilic property therefore affecting the enzymes involved in cell wall synthesis reactions, thus causes morphological changes in fungi which further lead to the lysis of the fungal cell wall [ 86 ].

4. Efficacy of Artemisia Oils against Insect Pests

Research has been conducted to see the effect of Artemisia oils against insect pests of agricultural crops, especially pests of stored products, in order to search out their efficacy as a repellent, insecticidal agent or antifeedant. From several national and international research institutions, investigators evaluated the essential oils from different species of genus Artemisia against storage and field insect pests. A. arborescens essential oil exhibited insecticidal effects against stored grain pest Rhyzopertha dominica at the dose of 50 μL in Petri dish [ 22 ]. A 37 μL/L and 24 h of exposure time of A. sieberi oil was sufficient to cause 100% mortality of Callosobruchus maculatus , Sitophilus oryzae and Tribolium castaneum. LC 50 (lethal concentration) values estimated for oil were 1.45 μL/L against C. maculatus , 3.86 μL/L against S. oryzae and 16.76 μL/L against T. castaneum [ 87 ]. In a filter-paper arena test, A. vulgaris oil had a very strong repellent activity against T. castaneum adults at a 0.6 μL/mL ( v / v ). In fumigation tests, 8.0 μL/mL dose of A. vulgaris oil exhibited 100% mortality of T. castaneum adults; mortality of larvae achieved was only 53%. A 20 μL/L air and a 96 h exposure of the oil showed 100% ovicidal activity; however, at a higher dose (60 μL/L) of this oil no larvae, pupae and adults were observed [ 88 ]. In fumigant toxicity test, 11.2 and 15.0 mg/L air LC 50 values were reported against Sitophilus zeamais adults, while in a contact toxicity test LD 50 (lethal dose) were 55.2 and 112.7 mg/adult for A. lavandulaefolia and A. sieversiana oils, respectively [ 89 ]. In another study [ 90 ], they found LC 50 5.31 and 7.35 mg/L, respectively for A. capillaris and A. mongolica essential oils against S. zeamais adults in fumigant bioassay, while in contact bioassay LD 50 values were 105.95 and 87.92 μg/adult, respectively. Again, A. scoparia essential oil achieved 100% mortality of C. maculatus at 37 μL/L air (24 h) in fumigant bioassay with LC 50 for the oil was 1.46 μL/L against C. maculatus and 2.05 μL/L air against S. oryzae and T. castaneum [ 91 ]. Similarly, 80–90% mortality of granary weevil, S. granarius (L.) was reported due to A. absinthium , A. santonicum and A. spicigera oils at a dose of 9 μL/L air after 48 h of exposure [ 92 ]. Against S. oryzae , A. princeps oil when mixed with Cinnamomum camphora , it showed strong repellent effect in 1:1 ratio and 1000 μg/mL of dose exhibited insecticidal action [ 93 ]. LC 50 value for A. vestita oil against S. zeamais in fumigant bioassay was 13.42 mg/L air, while LD 50 reported was 50.62 mg/adult in contact bioassay [ 94 ]. Later on, using same insect, they [ 8 ] determined 6.29 and 17.01 mg/L air LC 50 of A. giraldii and A. subdigitata oils in fumigant test and that of corresponding LD 50 40.51 and 76.34 μg/adult, in a contact test. EC 50 for A. annua oil was estimated to be 2.6 and 4.1 μL/mL against C. maculatus and T. castaneum , respectively [ 95 ], and LD 50 value of A. rupestris oil was 414.48 μg/cm 2 against Liposcelis bostrychophila and L. bostrychophila and 6.67 mg/L air LC 50 against L. bostrychophila [ 96 ]. This oil has also been proved as an effective insecticide against larval, pupal and adult stages of Helicoverpa armigera [ 97 ]. Plodia interpuntella , a polyphagous insect pest of different stored products worldwide, is found to be controlled by A. khorassanica essential oil (LC 50 : 9.6 μL/L air) with lethal time reported at 2.07 h [ 98 ]. Sharifian et al. [ 99 ] found that C. maculatus was more susceptible (LC 50 52.47 μL/L air) and T. castaneum was more tolerant (LC 50 279.86 μL/L air) towards A. vulgaris essential oil after 24 h of exposure. Respective LD 50 and LC 50 values of A. argyi essential oil determined by Zhang et al. [ 38 ] were 6.42 μg/adult and 8.04 mg/L air against Lasioderma serricorne adults. Their other report on A. stolonifera oil [ 45 ] showed LD 50 8.60 μg/adult against T. castaneum and 12.68 μg/adult against L. serricorne . The oil showed 1.86 mg/L air LC 50 value in fumigant test against T. castaneum . Liu et al. [ 36 ] reported that A. frigida essential oil exhibited 17.97 μg/adult and 254.38 μg/cm 2 LD 50 in contact toxicity test and 69.46 and 1.25 mg/L air LC 50 in fumigant test against adults of S. zeamais and L. bostrychophila , respectively. In contact toxicity, the corresponding LD 50 values of A. absinthium and A. herba - alba oils against T. castaneum , red flour beetle reported were 0.209 and 7.432 μL/L air [ 100 ]. In their further study with Oryzaephilus surinamensis LC 50 and LD 50 values of A. herba - alba and A. absinthium reported in fumigant and contact toxicity bioassay were 30.22 and 0.209 μL/L, respectively [ 100 ]. Recently, Liang et al. [ 101 ] reported insecticidal activity of A. anethoides oil by contact and fumigant tests against T. castaneum (LD 50 28.80 μg/adult and LC 50 13.05 mg/L air, resp.) and L. serricorne (LD 50 24.03 μg/adult and LC 50 8.04 mg/L air, resp.) adults.

Researchers also tested chemical constituents extracted from different species of Artemisia in order to make the botanical insecticides with a single and effective constituent. T rans -ethyl cinnamate (LD 50 0.37 μg/larva) isolated from A. judaica oil was more potent than piperitone (LD 50 0.68 μg/larva) against Spodoptera littoralis and also both these compounds caused complete inhibition of feeding activity at 1000 μg/mL [ 81 ]. 1,8-cineole and terpinen-4-ol (extracted from A. absinthium , A. santonicum and A. spicigera oils) were more effective against S. granaries with 100% mortality at 0.5, 0.75 and 1.0 μL/L air doses after 12 h of exposure [ 92 ]. Similarly, among chemical constituents of A. mongolica essential oil, 4-terpineol exhibited strongest contact toxicity (LD 50 8.62 μg/adult) against L. serricorne adults and camphor and alpha-terpineol in fumigant toxicity (LC 50 2.91 and 3.27 mg/L air, resp.) [ 41 ]. α-Terpinyl acetate (LD 50 92.59 μg/cm 2 ) of A. rupestris oil showed more contact toxicity than α-terpineol (140.30 μg/cm 2 ), 4-terpineol (211.35 μg/cm 2 ), and linalool (393.16 μg/cm 2 ) against book lice L. bostrychophila infesting stored cereals [ 96 ]. Some chemical constituents of A. argyi oil such as camphor (11.30 μg/adult), eucalyptol (15.58 μg/adult), β-caryophyllene (35.52 μg/adult) and β-pinene (65.55 μg/adult) exhibited more toxicity against L. serricorne adults having lower LD 50 values than that of α-terpinyl acetate, 4-terpineol, and linalool isolated from A. rupestris oil [ 37 ]. In their fumigant toxicity test eucalyptol (LC 50 5.18 mg/L air) and camphor (LC 50 2.91 mg/L air) had more toxicity than β-pinene (LC 50 29.03 mg/L air). Essential oil of A. ordosica possessed less toxicity (LC 50 18.65 mg/L air) against T. castaneum adults than its chemical constituents capillene, capillin, capillinol, cis -dehydromatricaria ester (LC 50 4.06 to 6.16 mg/L air) tested individually, however, among the essential oil and compounds tested, capillin showed strong repellency (100%) at 62.91, 12.58 and 2.52 μL/cm 2 after 2 h of exposure [ 102 ]. This revealed that the toxic properties of the oil could be attributed to the synergistic effects of its diverse major and minor components. All these results evidence that essential oils from these species of Artemisia oils and their constituents can be used in the formulation of botanical insecticides against the said insects for the long-term preservation of food commodities infested by these insects. The mechanism behind the insect mortality in the contact toxicity test is that the volatiles penetrate in the insect body via the respiratory system and result in abnormal breathing, which leads to asphyxiation and finally the death of insects [ 103 ]. During fumigant application, main target sites of essential oils and their constituents in insects is the octopaminergic system. When insects are exposed to the essential oils, a breakdown of the nervous system of insects occurs [ 104 ] which lead to the blockage of the nerve impulse, later paralysis and then death of the insects occurs.

5. Antioxidant Activity

Essential oils and chemical constituents of several Artemisia species have been investigated in the laboratory to protect against oxidative damage by inhibiting or quenching free radicals and reactive oxygen species. They have been proved as alternative antioxidants of synthetics. The antioxidant properties of the oils were assessed by several methods such as β-carotene bleaching (BCB) test, the 2,20-diphenyl-1-picrylhydrazyl (DPPH) radical scavenging method, thiobarbituric acid reactive species (TBARS), Trolox equivalent antioxidant capacity assay (TEAC I-III assay), Total radical-trapping antioxidant parameter assay (TRAP assay), N,N-dimethyl-p-phenylendiamine assay (DMPD assay), 2,2′-Azinobis 3-Ethyl-benzothiazoline-6-Sulphonate (ABTS), 2,2-diphenyl- l -picrylhydrazyl assay (DPPH assay), Photochemiluminescence assay (PCL assay) and Ferric reducing ability of plasma assay (FRAP assay) [ 105 ]. We assessed the antioxidant activity of A. nilagirica essential oil in our laboratory and found that the oil significantly inhibited radical cation formation, with 15.729 μL IC 50 (Inhibitory concentration) and 13.539 μL IC 50 preventing the bleaching of β-carotene [ 106 ]. While this oil exhibited higher antioxidant activity in the experiment of Sandip et al. [ 33 ] in the DPPH (IC 50 6.72 μg/mL) test, they reported the A. chamaemelifolia essential oil as weak antioxidant. Oil of A . scoparia induced secondary metabolites production in root cells viz., scavenging enzymes—superoxide dismutase, catalase, ascorbate and guaiacol peroxide and was phytotoxic to root growth causing its inhibition [ 107 ]. The A. annua essential oil (IC 50 27.07 mg/mL) was able to reduce the stable violet DPPH radical to the yellow DPPH-H, reaching 50% of reduction. However, IC 50 reported in ABTS method was 5.97 mg/mL lower than that of DPPH method. This oil was also 50% able to reduce the ferric ions to ferrous ions (Fe 2+ ) at 127.17 mg/mL [ 11 ]. This oil showed 18% antioxidant activity of the reference compound (tocopherol) [ 108 ]. Phenolic compounds present in the A. campestris essential oil contributed its major antioxidant activity, where 47.66 μg/mL EC 50 was reported in radical scavenging activity, 5.36 μg/mL in FRAP, 0.175 μg/mL in superoxide scavenging activity and 0.034 μg/mL in OH scavenging activity [ 109 ]. Thus, this oil can be used as an antioxidant in the pharmaceutical industry.

The pronounced antioxidant activity may be due to the phenolic constituents. A. campestris oil showed maximal DPPH activity at dose of 2 mg/mL [ 110 ], however, A. herba - alba oil showed strong DPPH activity (IC 50 6 μg/mL) than ABTS assay (IC 50 40 μg/mL) [ 72 ]. In another study, IC 50 values of A. turanica oil reported were 7.00 mg/mL, 9.69 μg and 14.63 μg, in DPPH, nitric oxide and superoxide anion radicals, respectively. The oil showed ferrous-ion chelating activity at 16.97 μg of IC 50 [ 59 ]. Ali et al. [ 111 ] reported that 0.005 mg/mL of ethyl acetate fraction of A. macrocephala oil showed 121.5% radicle scavenging activity. However, essential oil of A. deserti exhibited more antioxidant activity by DPPH free radical scavenging method (57.2%) than that of β-carotene bleaching test (50%) [ 17 ]. In the β-carotene method, A. dracunculus oil also showed 50% scavenging activity [ 21 ]. On the contrary, essential oils from A. absinthium , A. biennis , A. cana , A. dracunculus , A. frigida , A. longifolia and A. ludoviciana from Western Canada showed poor antioxidant activity in both the β-carotene/linoleate model and DPPH radical scavenging tests [ 112 ]. In addition, the antioxidant and DPPH radical scavenging activities of camphor and 1, 8-cineole isolated from Artemisia species were determined in vitro [ 113 ]. Singh et al. [ 114 ] reported more IC 50 (146.3 μg/mL) of A. scoparia than that of the antioxidant BHT (140.9 μg/mL) in DPPH bioassay. The residue essential oil also scavenged OH with an IC 50 of 145.2 μg/mL in the Fenton reaction using a deoxyribose assay. However, unlike scavenging of OH, residue essential oil exhibited a decreased scavenging activity towards H 2 O 2 (IC 50 270.1 μg/mL). They also reported that OH scavenging activities of citronellal and citronellol (25–200 μg/mL) were 8–34 and 11–55%, respectively. For the A. afra oil, 50% DPPH radicle scavenging inhibition was reported at 1.1 μL/mL, while it increased for A. abyssinica (28.9 μL/mL) oil. In lipid peroxidation bioassay only 0.09 μL/mL of oil is required for 50% inhibition [ 115 ]. From Tunisia, Riahi et al. [ 31 ] reported the variable IC 50 values (28.2 and 46.5 g/mL of leaf and flower oils, resp.) in A. absinthium oil. Additionally, essential oils from leaves (595.26 mol Fe 2+ /L) and flowers (286.42 mol Fe 2+ /L) also exhibited significant ferric-reducing antioxidant activity. From Serbia, A. annua oil showed 50% scavenging of radicle cations at 2.90 μg/mL in DPPH bioassay, and 50% antioxidant activity at 0.640 μg/mL in ABTS assay [ 24 ]. However, oil did not show superoxide-scavenging activity. IC 50 values for the chemical constituents in DPPH and ABTS methods reported were 4.00 and 1.79 μg/mL for Artemisia ketone, 87.0 and 30.1 μg/mL for α-pinene, 47.9 and 6.46 μg/mL for 1,8-Cineole, and 34.4 and 23.6 μg/mL for camphor, respectively. Mohammadi et al. [ 116 ] showed that A. absinthium essential oils extracted before flowering stage exhibited strong DPPH activity (EC 50 3.307 mg/mL) than that of the oils extracted at flowering (EC 50 4.11 mg/mL), and after flowering stage (EC 50 4.26 mg/mL). This may be due to presence of effective compounds such as sabinene, beta-pinene, alpha-phellandrene, p-cymene, and chamazulene which were more (58.36%) before flowering stage than that of at flowering (48.98%) and after flowering (53.99%). This may be also due to synergistic effect of the compounds [ 62 , 117 ].

6. Conclusions

Among herbal plants of the world, genus Artemisia biological activity is comparatively less explored against plant pathogens and insect pests. This genus is represented by more than 40 species, especially in the tropics. This review covers the chemical composition of essential oils from different geographical regions where a significant difference in the composition of different species of the same genus is observed. Major components consisted of several terpenes, terpenoids and phenolic compounds; and 1, 8-cineole, beta-pinene, thujone, artemisia ketone, camphor, caryophyllene, camphene and germacrene D were dominant in several species. The different Artemisia oils and their compounds have been reported as effective antimicrobial, insecticidal and antioxidant agents. Some oils also exhibited poor to moderate potency against pests and pathogens. Antioxidant activity found in oils is basically due to presence of phenolic compounds. The information summarized here is intended to serve as a reference tool to people in the field of plant protection and natural products chemistry. Although the current review focuses on the antimicrobial role of Artemisia essential oils against phytopathogens, it has also shown promising results against several human and animal pathogens. Recently this genus has attracted attention of the world when it commanded a Nobel prize regarding its use in traditional medicine for combating malaria. Although preliminary studies have been done on several species of Artemisia regarding its antimicrobial, antioxidant, insecticidal properties, elaborate bioprospection on its probable bioactivites against plant pathogens and pests is needed at field level. Recent times are desperate times where research interest has shifted towards exploration of natural compounds, especially for human welfare. More accurate reporting and data analysis is still needed. Other major issues such as mammalian toxicity, residual toxicity, phytoxicity and legal regulations/obligation and its long-term physiological and ecological effects of the effective oils need to be answered.

Conflicts of Interest

The authors declare no conflict of interest.

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Gut microbes associated with neurodegenerative disorders: a comprehensive review of the literature.

1. Introduction

2. materials and methods, 2.1. search strategy, 2.2. selection criteria, 2.3. data extraction, 2.4. data analysis, 3.1. type of study, 3.2. related disease, 3.3. methods of evaluation, 3.4. estimation of microbiome, 3.5. main findings, 3.5.1. parkinson’s disease, 3.5.2. alzheimer’s disease, 3.5.3. amyotrophic lateral sclerosis, 3.5.4. multiple system atrophy, 3.5.5. creutzfeldt–jakob disease, 3.5.6. huntington’s disease, 3.5.7. multiple sclerosis, 4. discussion, 5. limitations and future directions, 6. conclusions, supplementary materials, author contributions, data availability statement, conflicts of interest.

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Koutsokostas, C.; Merkouris, E.; Goulas, A.; Aidinopoulou, K.; Sini, N.; Dimaras, T.; Tsiptsios, D.; Mueller, C.; Nystazaki, M.; Tsamakis, K. Gut Microbes Associated with Neurodegenerative Disorders: A Comprehensive Review of the Literature. Microorganisms 2024 , 12 , 1735. https://doi.org/10.3390/microorganisms12081735

Koutsokostas C, Merkouris E, Goulas A, Aidinopoulou K, Sini N, Dimaras T, Tsiptsios D, Mueller C, Nystazaki M, Tsamakis K. Gut Microbes Associated with Neurodegenerative Disorders: A Comprehensive Review of the Literature. Microorganisms . 2024; 12(8):1735. https://doi.org/10.3390/microorganisms12081735

Koutsokostas, Christos, Ermis Merkouris, Apostolos Goulas, Konstantina Aidinopoulou, Niki Sini, Theofanis Dimaras, Dimitrios Tsiptsios, Christoph Mueller, Maria Nystazaki, and Konstantinos Tsamakis. 2024. "Gut Microbes Associated with Neurodegenerative Disorders: A Comprehensive Review of the Literature" Microorganisms 12, no. 8: 1735. https://doi.org/10.3390/microorganisms12081735

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