Remote Detection of Invasive Alien Species
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The spread of invasive alien species (IAS) is recognized as the most severe threat to biodiversity outside of climate change and anthropogenic habitat destruction. IAS negatively impact ecosystems, local economies, and residents. They are especially problematic because once established, they give rise to positive feedbacks, increasing the likelihood of further invasions and spread. The integration of remote sensing (RS) to the study of invasion, in addition to contributing to our understanding of invasion processes and impacts to biodiversity, has enabled managers to monitor invasions and predict the spread of IAS, thus supporting biodiversity conservation and management action. This chapter focuses on RS capabilities to detect and monitor invasive plant species across terrestrial, riparian, aquatic, and human-modified ecosystems. All of these environments have unique species assemblages and their own optimal methodology for effective detection and mapping, which we discuss in detail.
KeywordsInvasive alien species Biodiversity Hyperspectral remote sensing Lidar Resolution Unmanned aircraft systems (UAS) Unmanned aerial vehicles (UAV)
Invasive alien species (IAS) are non-native species with a rapid spread potential that can have negative ecological, environmental, and economic effects on the environments where they have been introduced (Masters and Norgrove 2010). The current rate and variety of species invasions is unprecedented in the fossil record (Ricciardi 2007). Global rates of invasion increased from around 8 records per year in 1800 to 1.5 per day in 1996. Although this rate may be partly the result of better record keeping, the rate is consistent across most taxa and shows little sign of slowing down (Seebens et al. 2017). Driven by climate change, invasion is expected to continue apace as global temperatures continue to rise and human societies and economies become increasingly connected around the world (Penk et al. 2016; van Kleunen et al. 2015).
12.1.1 Invasive Alien Species and Global Environmental Change
Human-mediated IAS introductions, deliberate or unintentional, tend to be much faster than natural processes (e.g., wind, animal; Theoharides and Dukes 2007; Hulme 2009; Pyšek et al. 2009; Seebens et al. 2017). Invasion pathways differ between taxa; intentional transport (escape and release) is most important for plants and vertebrates, while unintentional transport is more significant for invertebrates, algae, and microorganisms (Saul et al. 2017). Roads, tracks, and waterways create natural and artificial corridors for invasion, exposing ecosystems to invasion, particularly in emerging economies where development is rapid (Mortensen et al. 2009; Masters and Norgrove 2010). Globally, the continued expansion of tourism, air transport, and trade is dramatically heightening propagule pressure and subsequent invasion (Hulme 2015).
Global environmental changes, particularly changes in climate and weather patterns, nutrient cycles, and land use, generally drive increasing invasions while also making invasion prevalence, impacts, and feedbacks to the Earth system less predictable (Bradley et al. 2010; Dukes and Mooney 1999). These same change processes can also alter IAS transport and introduction mechanisms, hindering monitoring and control (Hellmann et al. 2008; Walther et al. 2009) and making it more challenging to predict future spread. Moreover, these changes stress ecosystems and increase invasion success (Simberloff 2000). Climate and land use changes drive species range shifts, potentially creating new invasion hotspots (Bellard et al. 2013; Bradley et al. 2010) while decreasing invasion risk and increasing recovery potential in other regions (Allen and Bradley 2016). Thus, observing the geographic patterns of the spread of IAS is critical to understand their origins, pathways, and invasion processes on a changing planet.
12.1.2 Biodiversity Impacts and Global Relevance
Biodiversity provides ecosystems with the capacity to respond to biotic and abiotic conditions and stress, often used as an indicator of ecosystem resilience. IAS threaten biodiversity through competition, hybridization, population reduction, and extinction of native species and modification of habitat. It has been estimated that 42% of all threatened or endangered species are at risk primarily because of IAS (Pimentel et al. 2005). IAS are able to thrive because they arrive in new ecosystems without coevolved local competitors, parasites, and pathogens to regulate their numbers (Keane and Crawley 2002) and are potentially able to exploit resources and niche spaces that natives cannot (Byers and Noonburg 2003; Levine 2000). Hybridization with local organisms reduces genetic diversity and further increases extinction risk (Mooney and Cleland 2001). For example, cheatgrass (Bromus tectorum) introduction to the Great Basin in North America resulted in decreases in biodiversity and dramatic changes in the ecosystem, as cheatgrass eliminated native competing shrubs (and thus species dependent on them) and increased fire frequency in the region (Pimentel et al. 2005). Ecosystem services losses and subsequent economic impacts of IAS are also high, from agriculture, forestry, and fisheries production losses to decreased recreation and tourism revenues (Pimentel et al. 2005). As of 2005, direct costs of invasive species and their management in the United States alone reach around $120 billion per year, excluding the degradation of invaluable ecosystem services (Pimentel et al. 2005). Globally, costs of invasions and IAS management exceed those of natural disasters by an order of magnitude (Ricciardi et al. 2011).
IAS-driven disturbances disproportionately affect developing countries, where livelihoods often depend on local natural resources that are threatened if IAS become prevalent (Masters and Norgrove 2010). Therefore, minimizing IAS spread is necessary to meet the targets in UN Sustainable Development Goal 15, Life on Land, which has a target focusing specifically on preventing introduction, controlling, and eradicating IAS. The European Environmental Agency (EEA 2012) has also developed an “invasive alien species in Europe” indicator summarizing the trends of invasions since 1900 and the greatest biodiversity threats. Meanwhile the US National Invasive Species Council coordinates and facilitates data interoperability across data providers and users, including defining data standards, formats, and protocols and facilitating cooperation across sectors and governments (National Invasive Species Council 2016).
In order to reduce the pressure of IAS on biodiversity and ecosystems, globally integrated approaches to IAS prioritization, management, and control are needed. Fundamental to international cooperation is cross-border policy and cooperation and transboundary assessments that are implemented within a global monitoring framework (Latombe et al. 2017). Following the Essential Biodiversity Variable (EBV) concept (see Fernández et al., Chap. 18), essential variables for invasion monitoring have recently been proposed by Latombe et al. (2017) to underpin a global monitoring system for IAS. Essential variables for IAS include occurrence, alien status, and alien species impact. Remote sensing (RS) is a valuable observation tool in this new EBV framework because it can be used to identify locations, cover, abundance, biomass, and other traits of IAS. Because it provides synoptic spatial, routine monitoring with fine scale, high-resolution RS can be used to identify sources of IAS and pathways for spread. RS-enabled IAS location data can inform control decisions and, with routine monitoring, can be used to quantify trends and predict invasion processes into the future to support policy decisions and management actions aimed at preventing undesired spread.
12.1.3 Remote Sensing for Detection of Plant Invasions
RS has long been favored as a tool for IAS mapping, specifically for plants, due to its ability to provide synoptic views over large geographical extents. This provides an advantage over field surveys, which are often limited to a small areas and may be in difficult to access locations. Historically, RS has been crucial in IAS detection. As far back as the 1970s, color infrared (IR) photos captured from airplanes were used to target herbicide applications to control water hyacinth (Eichhornia crassipes) infestations (Rouse et al. 1975). Over time, the state of the science has progressed substantially. Current technologies such as hyperspectral imaging spectroscopy and light detection and ranging (lidar) make it possible to detect and differentiate plant species within the same functional groups. Coupled with advances in image processing algorithms, these technologies have enabled accurate, repeatable RS measurements over time, providing consistent monitoring records to support control efforts.
Three factors make mapping IAS using RS most viable (He et al. 2015). First, when the IAS is the dominant growth form or has large homogeneous patches, it is easier to train a classifier to recognize it. For example, water hyacinth is sometimes the only IAS in lakes, so mapping it is as easy as separating bright green vegetation from spectrally dark water (Venugopal 2002). This is feasible with simple color IR aerial photography (Rouse et al. 1975) or multispectral satellite data such as that from Satellite Pour l’Observation de la Terre (SPOT) or Landsat. Second, when the target IAS has a unique phenology, it is easier to distinguish from native plants during some parts of the year. For example, Andrew and Ustin (2008) identified perennial pepperweed (Lepidium latifolium) during its flowering period, when it was spectrally most distinct from the surrounding marsh due to its unique white flowers. Temporally rich imagery can be used to identify the ideal time period for differentiation along with high spectral resolution to distinguish phenological differences. Third, the target IAS has a unique chemistry or biophysiology. For example, Khanna et al. (2011) differentiated water hyacinth from other co-occurring floating aquatic macrophytes using differences in canopy water content, since water hyacinth is a succulent with a higher plant-water content than co-occurring species water primrose (Ludwigia peploides) and water pennywort (Hydrocotyle ranunculoides). This requires a spectrally rich data set that is capable of quantifying canopy biochemistry. These three requirements are well matched with the three domains of RS data: spatial, temporal, and spectral.
Identify the target species and/or area. What IAS is affecting biodiversity, ecosystem services, or other economic functions in your area (e.g., transportation)? What do you know about your target IAS (e.g., spectral characteristics, phenology, ecosystem function, habitat requirements)? Do you know, or can you hypothesize, the IAS extent and community composition of other species in the area?
Determine the appropriate platform/sensor and identify/collect supplementary data based on species and habitat knowledge. Target species can be detected using direct or indirect methods. Direct detection uses spectral data and derived products from imagery. Indirect detection utilizes the ecological relationships between species and their environment to predict distribution.
Each species and habitat discussed in this chapter has specific characteristics that can be exploited to detect IAS. Exploitable differences can exist in the temporal, spatial, or spectral domains. The temporal domain consists of data collection timing and revisit timing. For example, if an IAS flowers at an earlier or later time than its surroundings, this information can be used to time image acquisition for when the target species appears most spectrally different. The spatial domain consists of pixel size and overall geographic coverage, and the spectral domain consists of the number of wavelengths, the position and bandwidth of wavelengths measured, and the spectral range of the sensor at which radiance can be measured reliably. Sensors typically have trade-offs among these domains based upon sensor design, size limitations, and data volume limitations. For example, in the spatial domain, there is a trade-off between overall coverage area or swath width and pixel size; both can be forced to increase, but at the expense of sensor size, which limits the platform it can be mounted on. There are also trade-offs between domains, mostly related to platforms. Most satellite platforms have larger pixel sizes than other platforms (20–100 s of meters) but have quick revisit time (days to weeks) and greater geographic coverage. Airborne platforms have a longer revisit time due to costs and logistics and smaller spatial coverage but offer smaller pixel size (centimeters to meters) and often support hyperspectral sensors. Unmanned aircraft systems (UAS) offer quick revisit time, on-demand deployment, and small pixel size but have very limited spatial coverage and limited spectral resolution due to size restrictions.
When direct detection is not possible due to canopy cover or other factors, indirect methods can be used to predict species locations. Species knowledge regarding habitat constraints or coexisting species can be used to govern a model using other data products. These data include things like digital elevation models (DEMs), climate layers, soil moisture, and any factor restricting species location. In some situations, these data can also be combined with direct detection methods to improve results.
Enhance data and model/classify. A model or classifier can be thought of as a set of rules or a mathematical function that uses pixel data to assign or predict class membership. This can either be supervised, where training data (pixels or spectra that have been identified previously) are used to define classes, or unsupervised, where classes are formed based upon pixel spectral/statistical similarity. Usually, atmospherically corrected surface reflectance data are provided to the classifier. Often, image enhancement is conducted to increase the information content of the input data. In addition to reflectance data, enhanced products can also be supplied to the classifier. Methods to enhance spectral data include spectral indices, principal component analysis (PCA), and minimum noise fraction (MNF). Spectral indices are combinations of spectral reflectance from two or more wavelengths that highlight a given reflectance or absorption feature and often indicate relative abundance of features of interest; for example, the Normalized Difference Vegetation Index (NDVI) is a normalized difference ratio of red and near-infrared (NIR) bands commonly used as an indicator of vegetation vigor.
With hyperspectral data, many narrowband indices are available that provide additional data about plant traits, including light use efficiency from the photochemical reflectance index (PRI; Gamon et al. 1997), canopy nitrogen from the normalized difference nitrogen index (NDNI; Serrano et al. 2002), canopy water content from the normalized difference water index (NDWI; Gao 1995), and a large number of leaf pigment indices [see Sims and Gamon (2002) for an overview]. Continuum removal is another technique used to target absorption features. For each pixel reflectance, a convex hull is fit over the top of the spectrum, absorption features are normalized to that hull, and the depth of a specific absorption feature (e.g., leaf water content) can be quantified. PCA is a linear transformation method that maximizes the variance of the data. When applied to a hyperspectral image, it produces a series of components that correspond to linear combinations of the original bands aligned to represent the variation within the original data set, with the first component being the plane responsible for the most variation. This allows for determining the most significant characteristics within an image that relate to classes. Minimum noise fraction transformation (MNF) rescales the noise in the data (a process called noise whitening), enabling the analyst to eliminate bands containing too much sensor noise and leaving only coherent image data.
Commonly used classification techniques include random forest, a supervised machine learning algorithm that constructs many decision trees and utilizes their outputs to get an accurate class prediction based upon training data, and maximum likelihood estimation (MLE), a supervised classification method in which parameter values of a statistical model are determined that maximize the chance that the process described by the model was actually observed. All of these data enhancement and classification methods can be performed using open-source software, such as R (https://www.r-project.org/) and Python (https://www.python.org/), where many packages are available to use, or in commercial software, such as ENVI (https://www.harrisgeospatial.com/).
Assess accuracy. One of the most important considerations is accuracy assessment following mapping. Depending on the objectives of the study, some types of error may be acceptable, while some may not. Typical accuracy metrics for image classification include overall accuracy, user’s accuracy, producer’s accuracy, and Kappa coefficient. Overall accuracy is the probability that an image classifier will correctly classify a pixel. This metric does not account for the number of validation pixels per class and may be misleading if a similar number is not used for each class. User’s accuracy and producer’s accuracy may be better metrics for assessing the classification. User’s accuracy (error of commission) is the fraction of correctly classified pixels with regard to all pixels classified. Producer’s accuracy (errors of omission) is the fraction of correctly classified pixels with regard to all ground reference validation pixels. In some situations, such as automated weed management in agriculture, overall accuracy and producer’s accuracy may not be as much of a concern as user’s accuracy because identifying small amounts of weeds (IAS) as crops may be okay, but spraying crops misidentified as IAS could be more damaging to crop yields than the IAS themselves. An example where maximizing producer’s accuracy may be more important would be in mapping IAS to understand species spread and the invasion process; any omitted species data as changes are monitored over time could affect process understanding and spread predictions. The last metric, the Kappa coefficient, can be useful for comparing multiple classification methods within the same data set. The Kappa coefficient is a measure of how closely the resulting overall accuracy of a classifier compares with expected accuracy, a random classification of pixels from the data set. One final consideration for accuracy assessment is the importance of having independent validation data that were not used in the mapping procedure. If accuracy is assessed with training data, it only measures how good the classifier is for those specific data that it was trained on, but the classifier may not be as accurate with other non-training pixels within the image.
12.2 Invasive Plants in Natural and Agroecosystems
Each ecosystem and IAS combination presents unique challenges for identification and mapping using RS. This is due to different landscape configurations, community composition, canopy structures, climates, habitat characteristics, and plant phenology. Each of these characteristics can be used to inform the optimal instrumentation for IAS detection and mapping. For this reason, we have separated IAS detection methods by biome and then split into more specific ecosystems and case studies.
Around one-third of Earth’s land surface is covered by forests. Forests are critical ecosystems, holding a very large proportion of global biodiversity. They are responsible for a large fraction of the global carbon storage and fluxes, strongly influence local and global water cycle processes, and provide fundamental goods and services to humanity (Foley et al. 2007). Globally, there are 26 types of forests, from taiga to tropical, all characterized by the unique ecological adaptations of trees to local climate, geology, and ecological conditions.
Forests invasions come in two types: (i) tree invasions (13 trees are in the top 100 world’s most invasive alien species, Lowe et al. 2000); and (ii) when other plants, such as vines and shrubs, or animals invade (Resasco et al. 2007; Cheng et al. 2007; Santos and Whitham 2010). Detection of invasion by tree species requires the direct detection of tree canopies (e.g., Asner et al. 2008a, b). Invasion of forests by other plants or animals can be detected directly, for example, when the IAS covers the canopy (Cheng et al. 2007), or indirectly, by measuring canopy leaf-off (Resasco et al. 2007; Wilfong et al. 2009), or through detection of pest impacts (Näsi et al. 2015; Ortiz et al. 2013).
Several studies have used optical RS data to directly detect invasion by tree species. One of the earliest approaches performed texture analysis on simulated satellite panchromatic imagery from historical 2 m aerial photography to map the invasive acacia (Acacia mearnsii) in South Africa (Hudak and Wessman 1998). Ramsey III et al. (2002) used 0.5 and 1.0 m color-infrared aerial photographs to map Chinese tallow (Sapium sebiferum) in Louisiana and Texas. They used a k-means classifier to discriminate IAS with relative success, attributed to the differences in senescence colors between the IAS and the native vegetation. A subsequent study scaled this approach to satellites, using a combination of Hyperion, Landsat 5, and aerial photos to define characteristic spectral signatures from 400 to 950 nm for Chinese tallow (Ramsey III et al. 2005). Pearlstine et al. (2005) also used aerial photos with larger spatial resolution (37 × 25 m) to map Brazilian pepper tree (Schinus terebinthifolius) using texture analysis on red, green, and NIR bands to identify the IAS relatively well.
Multispectral satellite data have been used to map tree IAS with varying levels of success. Fuller (2005) performed a supervised classification of IKONOS (2 m) and Landsat ETM+ (30 m) data to detect broad-leafed paperbark (Melaleuca quinquenervia) in Florida; the timing of imagery was chosen to enhance IAS separability. Cuneo et al. (2009) also used Landsat Enhanced Thematic Mapper (ETM) data to map African olive (Olea europaea cuspidata) in Australia based on spectral dissimilarity with the native Eucalyptus spp. with an accuracy of 85% and very low confusion between the species. More recently, decadal-scale time series afforded by sustained land imaging have enabled increased accuracy in cases where phenological cycles can distinguish IAS. Diao and Wang (2016) used a long time series to use the phenological changes in tamarisk for high-accuracy classification. Hoyos et al. (2010) mapped glossy privet (Ligustrum lucidum) in Argentina using a time series of Landsat TM data and machine learning (support vector machines, SVM), achieving classification accuracies of 89%.
Several studies used imaging spectroscopy to map tree IAS (He et al. 2011; Bradley 2014), e.g., tamarisk (Hamada et al. 2007; Carter et al. 2009), black cherry (Prunus serotina), black locust (Robinia pseudoacacia) and northern red oak (Quercus rubra) (Boschetti et al. 2007), Brazilian pepper (Lass and Prather 2004), and fire tree (Myrica faya) (Asner et al. 2008a, b). The studies determined characteristic IAS spectral profiles (sensu Ramsey III et al. 2005), compared spectral profiles across species using techniques such as SAM (e.g., Lass and Prather 2004), and correlated them with ground measurements (e.g., Asner et al. 2008a, b).
Direct detection of IAS on the tree canopy has also been studied. For example, Cheng et al. (2007) used imaging spectroscopy to detect kudzu (Pueraria montana) in a pine forest in Western Georgia, United States. They used an MNF transform and SAM to differentiate the spectral profile of the IAS from the native forest. Wu et al. (2006) mapped the invasive climbing fern (Lygodium microphyllum) in the Florida Everglades with a supervised classification of IKONOS imagery to show how it established in different parts of the forest. Although successful, their results underestimated fern extent in the understory.
Indeed, invasion of the forest understory is relatively understudied. Dense canopies mask understory contribution to the RS signal. To address this, researchers have turned to leveraging forest phenology to directly detect the understory when it is most visible. Resasco et al. (2007) mapped the historical spread of Amur honeysuckle (Lonicera maackii) during leaf-off conditions of the native forest using the Soil Adjusted Atmospheric Resistant Vegetation Index calculated from Landsat TM and ETM+ from 1999 to 2006. Wilfong et al. (2009) found that using a difference image measuring the difference between leaf-on and leaf-off conditions better predicted Amur honeysuckle cover than a single image. Evangelista et al. (2009) used a species distribution model to predict tamarisk (Tamarisk ramosissima) distribution over time based on vegetation indices derived from Landsat ETM+ data, with a 90% classification accuracy. Kimothi et al. (2010) used Indian RS satellite data to map another understory IAS, the West Indian lantana (Lantana camara), using texture analysis of images from September, February, and April. The dense leaf canopy meant separation of IAS was not possible in September, but classification accuracies were >90% in the other images after leaf fall. Barbosa et al. (2016) mapped subcanopy strawberry guava (Psidium cattleianum) outbreak with imaging spectroscopy and lidar and tested the accuracy of a machine learner, biased-SVM (BSVM), and mixture-tuned matched filtering (MTMF; a partial unmixing classification algorithm similar in principle to MNF) across canopy layers. While both methods allowed the estimation of the fraction of canopy layers that were invaded, the BSVM used information across the entire spectrum, while the MTMF did not, which may limit the applicability of MTMF when spectra of IAS are similar to “background” native species.
Indirect methods are another alternative to study understory IAS. Joshi et al. (2006) mapped Siam weed (Chromolaena odorata) in the understory using Landsat ETM+ and an artificial neural network to predict forest density and canopy light penetration and then subsequently predict Siam weed seed production. They found that 93% of the IAS seed production was predicted by the light intensity reaching the understory and concluded that this method worked relatively well to detect the IAS, despite the spatial resolution limiting detection to well-established IAS patches.
In summary, the most common method to detect tree IAS and map their distribution are to use their characteristic spectral signatures and dissimilarity with that of the native vegetation (Lass and Prather 2004). Tree IAS likely affect both the forest’s spatial structure as reflected in texture metrics (Pearlstine et al. 2005) and its 3-D structure, as shown with lidar (Asner et al. 2008a, b). To maximize the ability to detect invasive tree species, the use of the full visible (VIS) to shortwave infrared (SWIR) spectrum with imaging spectroscopy has shown clear advantages (Martin, Chap. 5), for example, in detecting the fire tree (Asner et al. 2008a, b) and for detecting bamboo (Dendrocalamus sp.) and slash pine (Pinus elliottii; Amaral et al. 2015). Alternatively, other studies selected specific bands that maximized discrimination and eliminated potential noise from nondiscriminating parts of the spectrum (Boschetti et al. 2007). While the advantages of imaging spectroscopy are obvious, data are not yet readily available to detect and map many tree IAS, especially in early stage invasion stages, although the upcoming launch of several hyperspectral satellite sensors will soon change this. Many tree IAS have different phenology than the native forest, either staying green longer, greening earlier, or flowering or budding later (Landmann et al. 2015); or they may be evergreen in a deciduous forest (Diao and Wang 2016). Timing imagery acquisition to maximize phenological differences has resulted in good classification accuracy (Ramsey III et al. 2002). Finally, using pixel sizes that match a tree canopy allows the detection of single invading trees (Bradley 2014). However, this can be very time-consuming, costly, and perhaps less systematic and viable across large areas or for early detection.
There are several analysis considerations for mapping tree IAS. A first, and perhaps most important, aspect is that tree IAS detection is prone to higher classification error (Bradley 2014) than other classifications, given the similarity in the spectral characteristics of trees to each other relative to other plant functional types. Spectral similarity between invasive and native trees may influence accuracy (Lass and Prather 2004), so ensemble classifications are recommended as well as other approaches that maximize spectral differences such as taking into account phenology. The examples cited above illustrate the value of a good field sampling design (Ramsey III et al. 2002) that covers the diversity of canopy structures (Hudak and Wessman 1998) and community compositions within the area of interest, since heterogeneity affects overall classification accuracy. In all of the studies discussed here, we observed a trade-off between omission and commission errors, where classification accuracy seems to be positively correlated with commission errors. Thus, we recommend that several accuracy metrics should be reported rather than just overall accuracy to give a better understanding of which species contribute to commission errors and which areas are more uncertain in IAS distribution maps.
12.2.2 Rangelands and Grasslands
Grasslands cover approximately one-third of the Earth’s surface (Latham et al. 2014), account for at least 30% of primary production by terrestrial vegetation (Grace et al. 2006), and, after forests, are the largest terrestrial carbon sinks (Anderson 1991; Derner and Schuman 2007; Grace et al. 2006). There are two main classes of grasslands, tropical/subtropical (also known as savanna) and temperate, which can further be described by three different subclasses: human generated, highly managed natural, and rangelands (Ali et al. 2016). Regardless of classification, these regions serve as a major source of animal feed and are heavily influenced by changes in climate and fire dynamics (D’Antonio and Vitousek 1992; Brooks et al. 2004). Contrary to popular belief, grasslands and rangelands harbor large amounts of biodiversity (Murphy et al. 2016); however, they are under threat as IAS continue to invade. This threatens biodiversity not only through direct losses by IAS replacing native grasses but also through indirect impacts to ecosystems by changing fire regimes (D’Antonio and Vitousek 1992; Balch et al. 2013), supporting wind erosion (Weisberg et al. 2017), and serving as a facilitator for plant viruses (Ingwell and Bosque-Pérez 2015).
Phenological differences are helpful for distinguishing native from non-native grasses. Given their frequent temporal resolution and global coverage, satellite optical sensors, such as Landsat TM/ETM+/OLI, SPOT, Sentinel-2, or, in some cases, Moderate Resolution Imaging Spectroradiometer (MODIS), have been used in several studies to map invaded grasslands. Cheatgrass, one of the top invaders in North America, greens up in early spring and senesces before native grasses, making it a suitable target species for RS approaches that leverage phenology differences (Fig. 12.4). Various studies across the United States have paired field data with multi-seasonal imagery selected during the green up (April–May) and senescent period to successfully map cheatgrass spread (Peterson 2005; Singh and Glenn 2009; West et al. 2017) and die-off (Boyte et al. 2015; Weisberg et al. 2017). Rather than just using images selected during green up and senescence periods, extracting phenology metrics from vegetation indices to refine cheatgrass classifications has also been successful (Bradley and Mustard 2008); however, in arid and semiarid environments, these indices can be highly influenced by rock and soil and should be used with caution (Singh and Glenn 2009). Huang and Geiger (2008) showed that a multi-date imaging approach can be successful even when natural phenologies of natives and nonnatives are similar. For example, native grasses and Lehmann lovegrass (Eragrostis lehmanniana) responded differently to unusual amounts of cool season precipitation, which allowed new tissues in invasive species to grow, making the two grasses distinguishable with multi-date imaging (Huang and Geiger 2008).
Imaging spectroscopy to map grassland IAS also often depends on differences in phenology, but the higher spectral resolution and typically higher spatial resolution afforded by airborne platforms often allow for more detailed and early detection maps. Image acquisition timing is important for species that exhibit differences in coloration throughout the year, such as flowering species or deciduous shrubs. In the case of leafy spurge (Euphorbia esula), hyperspectral instruments are better equipped to detect changes in flowering and thus have a higher success rate when compared to multispectral instruments (Mitchell and Glenn 2009). Leafy spurge has characteristic yellow flowers that bloom in early summer, and tamarisk leaves turn from yellow-orange to orange-brown in autumn before leaf drop. This distinct pigmentation enables remote detection using both imaging spectroscopy (Williams and Hunt Jr 2002; Glenn et al. 2005) and multispectral data (Anderson et al. 1993; Everitt et al. 1995; Evangelista et al. 2009). The blue-green color of new stems and the red-brown color of older stems help detection of spotted knapweed (Centaurea maculosa) from imaging spectroscopy (Lass et al. 2002; Lawrence et al. 2006). For early detection of goldenrod (Solidago altissima), an invasive moist tall grass in Japan, hyperspectral images acquired during early spring before full development of the grass canopy make it easier to map the exposed understory (Ishii and Washitani 2013).
Differences in canopy architecture or plant morphological traits, such as plant height and pubescence, can also be exploited when plants share similar phenologies or imagery is unavailable when growth cycles show key differences. Broom snakeweed (Gutierrezia sarothrae), for example, has an erect leaf canopy structure that results in a dark image response (Everitt et al. 1987; Yang and Everitt 2010). Spotted knapweed tends to inhibit the growth of other vegetation; the resulting increase in visible bare soil can help identify places where spotted knapweed grows (Lass et al. 2002; Lawrence et al. 2006).
In summary, multispectral sensors that provide free and open access to global imagery are used regularly for IAS detection in grasslands because their predefined temporal resolution offers recurring overpasses and at the very least provides seasonal imagery. This supports time series analyses and multi-date classification techniques. Looking to the future, changes in grassland species composition are anticipated to have the largest impact on Africa because it is home to the largest savannas, which cover roughly 50% of the continent (Campbell 1996; Grace et al. 2006). In addition, savannas in South American and Central Asia and temperate grasslands in the Western United States will also be heavily impacted. Thus, free, open-access, global mapping satellite RS data sets are especially important for grassland IAS detection and monitoring. These sensors lack the fine spatial resolution and full spectrum afforded by airborne imaging spectroscopy, which may be necessary to separate native and non-native species of the same functional type. However, multispectral imagery is often used for viewing widespread and abundant invasives, which is key for monitoring overall ecosystem invasion onset and die-off, but offers little help in terms of real-time or early IAS detection. In both cases, the minimum percent cover required for mapping can vary across similar ecosystems (Bradley 2014) and depends on sensor resolution and on how distinguishable the invader is from the background. Even when a non-native grass is spectrally distinguishable, an acceptable detection rate is not always possible when patch sizes are small relative to pixel resolution (Mladinich et al. 2006). Therefore, to ensure successful mapping, IAS targets must differ from the native community spectrally, phenologically, texture/morphologically, or architecturally (Bradley 2014). Analysis considerations must include a careful evaluation of the relationship between vegetation characteristics and sensor resolutions, particularly in the spatial, spectral, and temporal domains.
12.2.3 Aquatic Ecosystems
Although they cover a small portion of the Earth’s surface, aquatic ecosystems are disproportionately important to global diversity. They are among the most diverse and productive ecosystems on Earth and provide vital ecosystem services (Tabacchi et al. 1998; Barbier et al. 2011). Aquatic ecosystems encompass multiple gradients, such as water intermittency, microtopography, and salinity, leading to complex environmental heterogeneity (Junk et al. 1989; Mitsch and Gosselink 2007). This mosaic of diverse environmental conditions supports high biodiversity through multiple niches (Tockner et al. 2000; Ward et al. 2002).
Biodiversity losses in coastal and freshwater aquatic ecosystems are among the highest in the world (Dudgeon et al. 2015; Waycott et al. 2009; Vörösmarty et al. 2010). At least 30%–50% of the world’s wetlands have been lost (Finlayson 2012; Hu et al. 2017), and up to 35% of the extent of critical habitats like seagrasses and mangroves have been destroyed just in the twentieth century (UNESCO 2018). These ecosystems are among the most vulnerable to invasion because they are highly connected, are used extensively by humans, and often are geographically close to invasion foci such as ports or urban areas (Gherardi 2007; Williams and Grosholz 2008).
Plants in aquatic ecosystems can be broadly classified into five functional types or sets of species that occupy distinct spatial niches along the gradient from water to land and often have similar characteristics. The five functional types considered here are from land to water: riparian forests with shrubs and trees, emergent reeds and sedges, floating macrophytes, submerged macrophytes and macroalgae, and phytoplankton. Differentiating among these functional types with RS is achievable, but species-level detection within each community is more difficult due to similar survival strategies. Each functional type has its own challenges regarding species detection.
Many studies have successfully mapped IAS in aquatic environments using direct detection. Depending on the objectives of the study and the functional type being examined, spectral, spatial, and temporal requirements vary. In simple systems, high spatial resolution aerial photos can often be used to map species functional types as well as single species by taking advantage of unique attributes or phenology (Marshall and Lee 1994; Everitt et al. 1999, 2003). Mapping multiple species within the same functional types has been less successful using aerial photos. In these situations, more spectral information is needed to differentiate at the species level due to varying community complexity and species attributes (e.g., Khanna et al. 2011). Multispectral data have also been used with varying levels of success to map IAS in simple systems such as lakes invaded by just one species (essentially a two-class system; Venugopal 2002) or lakes with floating and some submerged vegetation (a three-class system; Everitt et al. 2003; Verma et al. 2003; Albright et al. 2004). Many classification methods have been used within aquatic ecosystems with varying degrees of success, including unsupervised classifiers, such as k-means and ISODATA (Ackleson and Klemas 1987; Dogan et al. 2009) and simple supervised classifiers, such as maximum likelihood and minimum distance (Malthus and George 1997; Vis et al. 2003; Nelson et al. 2006; Jollineau and Howarth 2008; Phinn et al. 2008; Yuan and Zhang 2008; Dogan et al. 2009), as well as more advanced machine learning methods (Malthus and George 1997; Nelson et al. 2006; Hestir et al. 2008, 2012; Everitt et al. 2011; Santos et al. 2012, 2016). While some studies have been successful and have even been operationalized into routine monitoring for invasive species management and reporting (sensu Santos et al. 2009; Santos et al. 2016), in many studies it is difficult to judge classification efficacy because accuracy assessment is missing or unusual, often not having independent validation data. Overall, machine learning algorithms seemed to have performed best. Within functional types, some specific strategies seem to work best as well. We highlight these below.
Riparian plants are often more difficult to differentiate at the species level than emergent and floating plants due to higher number of species and life forms, and a complex canopy structure, similar to forest IAS detection. Riparian IAS sometimes grow in monocultures, which may be easier to detect (e.g., giant reed, Arundo donax). Other IAS can grow embedded in the native community similar to grasslands, making them harder to map using RS (e.g., yellow star-thistle, Centaurea solstitialis). From an RS perspective, the layered canopy, many species, and mixed pixels make it hard to map target IAS within this complex community mosaic.
Community complexity can often be overcome by taking advantage of differences in phenology. Acquiring imagery during flowering or senescence when the target IAS is most distinct from its surrounding vegetation may allow for detection at the species level. For example, Landsat ETM+ and QuickBird have been used to take advantage of correct timing and fine spatial resolution, respectively, to distinguish riparian IAS (Laba et al. 2008; West et al. 2017). Frequently, increasing spectral data further has been necessary to detect riparian IAS. (Ustin et al. 2002; Laba et al. 2005; Hamada et al. 2007; Andrew and Ustin 2008).
Within the emergent functional type, the canopy is relatively uniform, composed generally of only grasses, sedges, and reeds. These species are often mixed, and patch sizes remain small even among species tending to grow as monocultures. The canopy structure is typically erectrophilic, and spectral mixing with water is common, even with fine spatial scale imagery. In addition to spectral information and temporal information, the texture of invasive and native species patches can be leveraged in mapping IAS and may be used to improve success. Samiappan et al. (2017) used four methods to calculate texture indices as inputs into a SVM algorithm to map common reed (Phragmites australis). They took advantage of the 5 m spatial resolution afforded by airborne (in this case UAS) imagery, though they cautioned such an approach is unlikely to work if patches of IAS are smaller than a few pixels or more mixed. However, texture has been shown to be advantageous even with moderate spatial resolution imagery. For example, Arzandeh and Wang (2003) successfully differentiated common reed and cattail (Typha angustifolia) using Landsat TM by adding texture indices to increase pixel spectral information content. For these reasons, hyperspectral aerial surveys have offered the best data source for classifications for emergent communities. Using sensors such as CASI, AVIRIS, and HyMap, many studies have mapped the emergent community, differentiating species within submerged and floating functional types (Hestir et al. 2008; Jollineau and Howarth 2008; Hunter et al. 2010; Khanna et al. 2011; Hestir et al. 2012; Zhao et al. 2012). Occasionally, both spectrally rich and temporally strategic data have been used together to map IAS (Laba et al. 2005; Hamada et al. 2007; Pu et al. 2008).
126.96.36.199 Floating Macrophytes
188.8.131.52 Submerged Macrophytes
Mapping submerged macrophytes and macroalgae presents additional challenges due to the presence of the water column. Detection of these plants is complicated by the combined effects of inherent optical properties (IOPs) of the water column, which are influenced by the diffuse attenuation of the water column itself and the absorbing and scattering properties of its dissolved and suspended matter, and the apparent optical properties (AOPs), which are controlled by weather, sun, and sensor view angles (which can lead to sun glint or insufficient signal returns) as well as the influence of the air-water interface (Mertes et al. 1993; Bostater Jr. et al. 2004; Morel and Bélanger 2006; Hestir et al. 2008). IOPs are difficult to account for because water quality and depth can vary spatially and temporally with runoff, geomorphological gradients, meteorological conditions, flow conditions, land use practices, tidal stage, and phytoplankton phenology and community changes (Vis et al. 2003; Nelson et al. 2006; Hestir et al. 2008). Radiative transfer approaches are useful for classifying submerged species. Typically, they use either use model inversion or look-up tables to solve the radiative transfer model in the water column to distinguish different optically active constituents (e.g., phytoplankton and other pigments, suspended non-algal particulates, colored dissolved organic matter) and differentiate bathymetry and bottom type (see Odermatt et al. 2012; Giardino et al. 2018 for comprehensive reviews of the approach). These approaches are often preferred because, being physics-based, they are in principle generalizable and transferable across sensors and systems (Giardino et al. 2010, 2012; Malthus et al. 2012; Hestir et al. 2015). However, such approaches require detailed spectral information on specific water body IOPs, which are difficult to collect and not generally available (Matthews 2011; Lymburner et al. 2016). In these approaches, bottom type is typically mapped to just a few broad classes (e.g., sand/sediment, rock, submerged plants, coral), so species-level detections are not common in the literature (Dörnhöfer and Oppelt 2016). However, Santos et al. (2012) were able to show species-level discrimination of submerged macrophytes at the leaf level and could differentiate native from non-native submerged macrophytes at the canopy level from HyMap airborne imaging spectroscopy in a turbid estuary in California.
To circumvent some of the confounding factors of the air-water interface and water column for mapping submerged macrophytes, hydroacoustics are often used for bed delineation and height and density quantification (Winfield et al. 2007). These require intensive boat surveys (which limit access), do not provide species-level discrimination, and can provide significantly different results for the same system due to lack of standardization in signal processing approaches (Radomski and Holbrook 2015). Recently it has been argued that RS imagery approaches are, despite several limitations, overall more efficacious than hydroacoustic surveys (McIntyre et al. 2018).
Commercial shipping and the exchange of ballast water is one of the main pathways of IAS spread in marine and aquatic environments around the world. It is difficult to characterize phytoplankton species as native or non-native due to limited inventories, varying morphology and complex synonymy based on regional environmental differences, and the spontaneous “appearance” of new species (Olenina et al. 2010). Nonetheless, many phytoplankton species have been documented to have spread via ballast water (Subba Rao et al. 1994; Olenin et al. 2000), and species recorded in ships’ ballast water are increasing in abundance (Olenina et al. 2010). Rapid shifts in species composition and large harmful algal blooms in coastal and inland waters have cascading effects on community structure for waterfowl, marine mammals, fish, shellfish, and benthic communities and are a constant concern for biodiversity conservation and ecosystem managers (Anderson et al. 2002).
In the water column, different phytoplankton pigments have key spectral absorption features that can be resolved in order to make inferences about their functional type. Chlorophyll a, the key diagnostic pigment for many diatoms, absorbs strongly at 435–438 and 660 nm. Cyanobacteria, the common culprit of large-scale harmful “blue-green” algal blooms, show absorption features at 490–625 nm. Floating algae have spectral features in the 550–900 nm range. Mesodinium rubrum, the photosynthetic ciliate that causes red tides, contains the pigment phycoerythrin, which fluoresces in the yellow peak (565–570 nm; Dierssen et al. 2015).
With the exception of key diagnostic pigments that allow direct estimation of the concentration of certain species (e.g., coccolithophores, Mesodinium), RS of phytoplankton species is typically limited to detection of phytoplankton functional types or groups (based on taxonomic criteria or biogeochemical function) or phytoplankton size class (based on size range) (Bracher et al. 2017). Most detection algorithms rely on radiative transfer models that account for bio-optical properties (e.g., pigment composition, absorption, and backscattering), empirical relationships that relate chlorophyll a concentrations measured via satellite with in-situ measurements of diagnostic marker pigments determined from high-performance liquid chromatography (HPLC) or ecological models that predict phytoplankton functional type presence based on different abiotic and biotic parameters. Moisan et al. (2012) and Bracher et al. (2017) provide an overview on the state of the science for RS phytoplankton species detection. Sathyendranath et al. (2014) and Mouw et al. (2017) provide details on most of the current algorithms and procedures for phytoplankton functional type mapping from RS.
Mapping phytoplankton functional types in coastal and inland waters is still challenging, however. Current land missions lack the temporal resolution to make frequent, repeated observations at the scale of tidal, riverine, meteorological, and biotic processes (e.g., growth, grazing, senescence) that drive phytoplankton variability (Muller-Karger et al. 2018). Phytoplankton and water quality change on the scale of hours to days due to runoff, advection, and mixing. Kudela et al. (2015) used time series of field hyperspectral observations to show that phytoplankton blooms can be displaced by cyanobacteria in a few days. Hestir et al. (2015) documented similar rapid changes in cyanobacteria from hyperspectral measurements. Chen et al. (2010) observed phytoplankton blooms that evolve over 2–3 days in Tampa Bay. After 13 years of observations in Long Island Sound, Dierssen et al. (2015) concluded that monthly measurements are insufficient to quantify episodic plankton blooms. While they documented a bloom of a ciliate that could only be detected with hyperspectral measurements, of yellow fluorescence, only one such image has ever been collected of this area and this was with the Hyperspectral Imager for the Coastal Ocean (HICO) that ceased operations in 2014.
In summary, RS of aquatic IAS requires moderate to fine spatial resolution, high spectral resolution, and, for submerged IAS, high radiometric resolution. We are optimistic that future global mapping missions with climate-relevant mission durations can improve riparian and aquatic IAS mapping by enabling time-based approaches such as phenology signatures. Even without high spectral resolution data, RS of aquatic macrophytes is progressing. For example, through radiative transfer modeling, it has been shown to be robust for mapping aquatic macrophyte morphological traits in temperate systems (e.g., leaf area index, fractional cover, and biomass) across floating, emergent, and submerged macrophytes, which can be used to better quantify nutrient uptake, community dynamics, and invasion hotspots (Villa et al. 2014, 2015, 2017). The rapidly developing science of drone and UAS imagery also raises the potential to map IAS using differences in texture or using segmentation tools to do object-based mapping, especially when the area being mapped is small.
Agroecosystems are unique ecosystems due to the extraordinarily high anthropogenic interventions and pressures placed on them. Unlike other ecological systems, agricultural systems have more controlled environmental conditions with limited plant biodiversity. Crops are often grown as a monoculture, in uniform rows with highly regulated demography. Though crop species are often robust and herbicide resistant, many IAS are also developing resistance to herbicide, making them more invasive with increasing impacts on crops. With a rising global population, there is increased pressure on agricultural systems to increase productivity. IAS consume resources meant for crops and reduce yield, productivity, and income for farmers. In corn and soybean, two of the major crops grown in the United States losses due to IAS have been estimated at $17 billion in soybean and $27 billion in corn annually, approximately 50% of the yield of each of these crops (Soltani et al. 2016, 2017). IAS can become established in agroecosystems as in any other system, through both natural (wind, water, animals, forceful dehiscence) and artificial (machinery, crop seed, livestock feed, spreading of crop, and livestock waste) means. The application of water and nutrients also complicates the system by enhancing IAS’ ability to compete with crops and reproduce. Often the effects of IAS depend on the crops present. Certain IAS may be problematic in some crops but not others due to crop management practices (time of planting, tillage, irrigation, mulch, registered herbicides, rotation).
To effectively detect IAS in agricultural systems, RS must meet the challenge of detecting IAS before they become competitive with crops. Field spectroscopy has been shown to be effective for discrimination of IAS from crops (Basinger 2018; Koger et al. 2004a, b; Gray et al. 2009), but it is not the most efficient due to the short duration of such field campaigns, since detection must then occur within a small window during one growing cycle. Research has long been published on the use of satellites or other airborne sensors for IAS detection in agriculture (Hunt et al. 2007; Menges et al. 1985), but these methods often lack the spatial and/or temporal resolution needed to detect IAS intermixed with a crop species.
One factor that aids in detection is that IAS tend to emerge in patches or patterns associated with farm management practices. For example, plants growing outside of the uniform row formations are often IAS and can be treated. Most studies typically investigate a single IAS. However, IAS are often intermixed, making them hard to distinguish from each other. Additionally, Basinger (2018) found that using field spectroscopy, IAS detection is not uniform across cropping systems and suggested that improved IAS detection may require crop-specific parameters for accurate IAS detection and control.
The main challenges of using RS in agroecosystems are associated with data latency (which impedes rapid IAS management on the part of producers) and the necessity of early growth cycle detection (where many species appear similar). Current market solutions tend to focus on active sensors or the use of artificial lighting rather than passive sensors. Commercial early IAS management systems used active proximal sensors to spot and spray IAS with herbicides. However, while these systems can detect vegetation, they are not able to detect small IAS (Blackshaw et al. 1998) or distinguish between the crop and IAS. They thus rely solely on a priori assumptions about timing of emergence of IAS relative to crop species.
In summary, using RS in agroecosystems is only useful to growers within the timeline of crop cycles. IAS control is most effective when plants are small, but this is when they are also most difficult to detect and differentiate from the crop. IAS detection often requires high spatial and temporal resolution due to synchronous and asynchronous IAS emergence with the crop and sometimes high spectral resolution to deal with similar appearances during early growth stages. Implementing data-based management decisions is difficult if monitoring is not near constant due to the necessity of rapid responses. Thus, the use RS for the control of IAS has seen limited adoption in agriculture, despite a long history of research. However, UAS have become more common because the technology now meets several of the requirements for RS of IAS in agricultural settings.
12.2.5 Urban Ecosystems
More than half of all people live in urban areas, and this proportion is expected to increase substantially during this century. Urban ecosystems differ from agricultural or natural systems in terms of structural properties related to the built/natural ratio of the landscape; built area includes impervious and permeable built environments and the connecting infrastructure. Urban ecosystems have been colonized by increasing numbers of IAS (Paap et al. 2017; Hui et al. 2017). These ecosystems are unique because trees and other ornamental species in private and public city gardens are often non-native and can be sources of IAS to surrounding areas (Paap et al. 2017; Mayer-Pinto et al. 2017). IAS richness in urban areas is positively correlated with housing density (Gavier-Pizarro et al. 2010), urban wastelands (Bonthoux et al. 2014; Maurel et al. 2010), green infrastructure (Hostetler et al. 2011), and roads (Rupprecht et al. 2015). By harboring IAS, cities may unwittingly act as sources of IAS to surrounding agroecosystems and natural ecosystems (Paap et al. 2017; McLean et al. 2017).
Use of RS for IAS detection and mapping in urban environments is essential to gauge the affect of urban plants, which are often non-native, on the surrounding ecosystems. Detection has been successful with many forms of RS. For example, Shouse et al. (2012) used a combination of 0.3 m color aerial photographs and multispectral Landsat data to map bush honeysuckle (Lonicera maackii) under the forest canopy in an urban park in Louisville, Kentucky, USA. They conducted an object-based classification, a supervised classification, and constructed a species distribution model, with accuracies above 75%, especially for the object-based classification. This high accuracy can be attributed to extended greened-up seasons and high spatial resolution. Hyperspectral data has been used to detect Himalayan blackberry (Rubus armeniacus) and English ivy (Hedera helix) in nonforested areas of Surrey, British Columbia, Canada (Chance et al. 2016a). Classification accuracies were higher than 75% for both; the potential for spectral separability was maximized by the choice of wavelength regions, and the researchers were able to increase accuracy using a random forest classifier, due to higher capability of under-canopy detection (Chance et al. 2016b). Because urban ecosystems are smaller than other ecosystems and more complex, high spatial resolution is necessary to detect IAS within the mostly non-vegetative ground cover.
Lidar with spectral data also has proven effective for mapping vegetation within urban areas. By combining lidar with hyperspectral imaging and a random forest classifier to map tree species including honey locust (Gleditsia triacanthos) in Surrey, British Columbia, Canada. Liu et al. (2017) further improved classification accuracy, showing the power of data fusion. Other studies have combined lidar data with IKONOS multispectral data to detect whether Chinese privet (Ligustrum sinense) invasion changed urban forest structure in Charlotte, North Carolina, USA (Singh et al. 2015). These researchers also found that a random forest built with lidar-derived metrics produced the best results.
RS of urban IAS, however, has some unique challenges. Because most of the ground is covered by manmade features, it is difficult to detect green areas and map and identify individual species (Alonzo et al. 2014). With sufficient spatial resolution, these challenges can be overcome. The most successful approach to date is to use a combination of hyperspectral and lidar, which yields spectral, structural, and height information.
In summary, detection of IAS in urban environments requires high spatial resolution to differentiate natural from built environments, high spectral resolution to identify species, and sufficient temporal resolution to detect IAS at different stages of invasion. While this is an emerging field with a growing literature, relatively few studies of IAS in urban environments have used RS data, and further research is needed in different geographical settings, invasion process phases, and urban density conditions.
12.3 Summary, Conclusions, and Prospectus
Invasive species are a major direct driver of biodiversity loss because they outcompete native species for local resources, eventually replacing or displacing them. They also cause indirect losses because they do not assume all of the ecological roles of the replaced native species. As they spread, IAS modify nutrient availability, nutrient cycling, soil chemistry, water quality, hydrology, food webs, habitats, and other ecosystem functions (Gordon 1998; Scheffer et al. 2003; Dukes and Mooney 2004; Hestir et al. 2016; Khanna et al. 2018), impairing ecosystem function. In addition to causing functional changes, IAS also modify ecosystem structure by physically changing canopy structures in forests and water quality in aquatic ecosystems. Increasing global changes related to climate, nutrient cycles, and land use will potentially change transport and introduction mechanisms of IAS in a way that provides a competitive advantage for new IAS, likely reducing effectiveness of control strategies. The acceleration in global change and biodiversity loss degrades ecosystem resilience, threatening valuable ecosystem services. To preserve these services will require global cooperation on IAS monitoring and control with RS is a critical tool.
Each biome discussed in this chapter contains a unique complement of species. As a result, a different method of RS and data fusion works best for each. However, some methodologies can be valuable in all circumstances, such as increasing spectral information content. If only one IAS is of interest and it differs from its surroundings, multispectral data or use of photographs and texture analysis may be enough to identify and map it. However, in most cases there are multiple IAS competing with one another and with native vegetation, with varying canopy complexity and functional types. In such scenarios, difference in phenological characteristics can be exploited for identification. For example, an IAS might be identified through differing flowering times, flower colors, or earlier or later periods of senescence relative to surrounding vegetation. This requires temporally dense data. In cases where the invasion scenario is not simple or the data are not temporally sufficient, fusion between RS and other data sources (e.g., habitat models, DEMs, climate models) can be used to improve accuracy.
Data collection in the three domains of RS (spectral, spatial, and temporal) can be optimized for a species based on the ecosystem type and image analysis approach. For forests, lidar data are often a good addition to spectral information because they can provide information on height and physical crown structure. For species below the forest canopy, indirect methods such as models based on ecological knowledge of the species may be necessary, or imagery may simply be collected during a leaf-off period. IAS in grasslands often have similar spectral properties to natives, requiring hyperspectral data, strategic image timing, or indirect modeling methods. Aquatic ecosystems introduce many confounding factors due to presence of water and its associated processes, necessitating high radiometric quality and good calibration. Because this biome is so complex, hyperspectral information and customized image timing are a must for differentiating IAS. Additionally, radiative transfer modeling is often necessary to detect submerged and water column IAS. Agroecosystems have minimal diversity, so fewer spectral data are required. However, frequent assessment is necessary to allow a timely response to minimize crop loss. RS detection of IAS in urban ecosystems requires varying methods and unique adaptations because of the high potential for introductions and unusual landscape features, such as impervious surfaces.
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