Tag: ERAdiversity

The following articles lay out a few key ecological concepts and terms that may be helpful to become familiar with for the growing number of biodiversity-conscious people and organizations that are beginning to plant more native plants on their land.

Native plants, native ecosystems, and native landscapes: an ecological definition of “native” will promote effective conservation and restoration, Wilson, Hibbs & Alverson 1991 

Produced by the Native Plant Society of Oregon, this article argues that, while the use of native species is an accepted tenet of conservation, the term “native” is not necessarily well understood; they attempt to clarify the term.

“Any definition of a native species, native ecosystem, or native landscape requires an historical benchmark” [Wilson 1991: 16]. Over the past 20,000 years, “vegetation in the Willamette Valley has changed dramatically with changing climate. Vegetation in a single place has probably varied from boreal parkland, to conifer forest, to oak savanna, to prairie. Each climatic phase supported a different flora” [Wilson 1991: 16]. Each of these vegetation types was native to a particular place, according to particular climatic conditions that changed overtime. The vegetation that developed in the past 10,000 years – the current Holocene period of climate stability – is thus the relevant reference.

“For the Pacific Northwest, the period that ended with Euro-American settlement is a natural historical benchmark. This period lasted long enough to have a significant impact on the vegetation of the region. The climates of much earlier times were different enough to limit their usefulness in defining today’s ecosystems” [Wilson 1991: 16]. Thus, “any species that had occurred in a particular ecological habitat [of the Pacific Northwest] before Euro-American settlement is a species native to that habitat” [Wilson 1991: 17].

A native ecosystem, then, is one dominated by native plants, animals and microorganisms that occurred together before the time of Euro-American settlement. Key species – for example, the dominant photosynthesizing plants, the top carnivores, the important decomposers, the nitrogen-fixers – must be present for a native ecosystem to persist and function on its own. To artificially maintain a conserved or restored ecosystem without all of its crucial components is both difficult and expensive. The species of native ecosystem must also occur together in nature. For example, landscaping with an artificial mixture of native species like vine maple, blue bunch wheatgrass, and Jeffrey pine does not produce a native ecosystem. These species are native to different areas within Oregon, but they would not naturally grow together in the same ecosystem. Restoration of native ecosystems must also account for proper structure and appearance. For example, a red fescue lawn does not have the structural complexity and species diversity exhibited by native bunchgrass prairies [Wilson 1991: 17].

Key species – for example, the dominant photosynthesizing plants, the top carnivores, the important decomposers, the nitrogen-fixers – must be present for a native ecosystem to persist and function on its own. To artificially maintain a conserved or restored ecosystem without all of its crucial components is both difficult and expensive [Wilson 1991: 17].

The community as an ecological unit, Barbour, Burk & Pitts 1987

This article provides an overview of types of plant communities and the process of succession in those communities.

In each type of habitat, certain species group together as a community. Fossil records indicate that some of these groups (or very closely related precursors) have lived together for thousands or even millions of years. During that time, it is possible that an intricate balance has been fashioned. Community members share incoming solar radiation, soil water, and nutrients to produce a constant biomass; they recycle nutrients from the soil to living tissue and back again; and they alternate with each other in time and space. Synecologists attempt to determine what is involved in this balance between all the species of a community and their environment [Barbour 1987: 155].

Community concepts and attributes

A plant community is an identifiable stand of plants growing together in a certain spot. Clusters of species, called associations, are often found growing together in several different places within a larger region. “An association is a particular type of community, which has been described sufficiently and repeatedly in several locations such that we can conclude that it has: (a) a relatively consistent floristic composition, (b) a uniform physiognomy [appearance], and (c) a distribution that is characteristic of a particular habitat” [Barbour 1987: 156].

There are opposing views about why particular plant species are often found growing together in a plant community. The continuum view posits that species distribution is driven individualistically by each species’ particular tolerance to various environmental conditions. By contrast, the association view suggests that a plant community is an integrated whole, whose component species are interdependent.

Whatever the reasons that particular species tend to grow together in stands, however, such stands “exhibit collective or emergent attributes beyond those of the individual populations” [Barbour 1987: 159]. Examples of such community attributes include its vertical structure, canopy cover, species composition and diversity, biomass, productivity, stability, and nutrient cycling, for example.

Succession

Ecological succession is an important concept that helps explain the particular assemblage of plants growing in a given location.

“Plant succession is a directional, cumulative change in the species that occupy a given area through time” [Barbour 1987: 230]. This does not refer to cyclical changes that occur over seasons, nor to changes occurring in response to climate shifts over extremely long time spans like thousands or millions of years. Rather, succession is when the composition of plants at a particular site changes over a period of decades to centuries.

Succession begins when pioneer species colonize bare ground. These first arrivals tend to be opportunists that grow fast, reproduce quickly, and do not live long. The early successional plants start to improve the habitat conditions for other, more competitive plants to then take over, displacing the pioneers. “One of the driving forces behind succession is the effect plants may have on their habitat. Plants cast shade, add to the litter, dampen temperature oscillations, and increase the humidity, and their roots change the soil structure and chemistry. … Both the environment and the community change, and this metamorphosis is due to the activities of the organisms themselves.” [Barbour 1987: 233]

Overtime, slower-growing, larger, longer-living plant species outcompete the earlier successional species, eventually forming a climax community, which is not subsequently replaced by any other community. “Succession often leads to communities with greater and greater complexity and biomass and to habitats that are progressively more and more mesic (moist)” [Barbour 1987: 233]. Such changes result in climax communities tending to be self-sustaining due to efficient nutrient cycling and internal moderation of external fluctuations in temperature and humidity.

The particular plant composition of a climax community depends on the regional climate, as well as local soil conditions and topography, meaning that several climax communities can exist in a given landscape.

Typically, many plant communities coexist in a complex mosaic pattern. That is, one climax community does not cover an entire region. … In [some] cases, the mosaic reflects topographic differences, such as south-facing versus north-facing slopes, basins with poor drainage and fine-textured soil versus upland slopes with good drainage and coarser soil, or different distances from a stress such as salt spray. In such cases, the communities within the mosaic do not bear a successional relationship to one another; they constitute a toposequence. Each community in a toposequence may, in fact, be a climax community [Barbour 1987: 238].

Understanding ecological succession can help us to predict the future vegetation of a site by observing its current vegetation. “It is often possible to estimate a community’s future composition by extrapolation from changes measured in a short time, by comparing other communities that have plants of different ages, or by noting differences between overstory plants and understory seedlings” [Barbour 1987: 231] In some cases, the understory seedlings will later become the canopy, provided the localized conditions support this succession.

Vegetation Ecology: Historical Notes and Outline, van der Maarel & Franklin 2013

These authors define the concept of a plant community through discussion of its evolution. They start by defining the term ‘vegetation’ in a way that may surprise some readers because it excludes plants growing in certain situations. To be considered vegetation, plants need to emerge spontaneously.

Vegetation, the central object of study in vegetation ecology, can be loosely defined as a system of largely spontaneously growing plants. Not all growing plants form vegetation, for instance, a sown corn field or a flower bed in a garden do not. But the weeds surrounding such plants do form vegetation. A pine plantation will become vegetation after some years of spontaneous growth of the pine trees and the subsequent development of an understory [van der Maarel 2013: 1].

Two competing schools of thought regarding the nature of a stand of plants growing together geographically are represented by two early 20th Century botanists. H.A. Gleason observed “that species are ‘individualistically’ distributed along omnipresent environmental gradients and thus cannot form bounded communities” [van der Maarel 2013: 2]. By contrast, E. Clements compared plant community with an integral organism, where the whole was greater than the sum of its parts. During the same time period, the Braun-Blanquet approach was developed, which “paid much attention to the relations of plant communities with the environment and the interactions within communities, which is now incorporated in the concept of ecosystem” [van der Maarel 2013: 2].

The authors state that while individual plant species are distributed according to abiotic environmental conditions, the fact of being co-located with particular sets of other species in a particular environment results in interspecies interactions, which are in fact ecosystem processes (emergent properties).

In conclusion, a plant community is generally recognized as a relatively uniform piece of vegetation in a uniform environment, with a recognizable floristic composition and structure, that is relatively distinct from the surrounding vegetation. Even if the populations of the participating species are usually distributed individualistically in the landscape, they may well interact within the community and build up an integrated unit with emergent properties. At the same time, plant communities can be convenient units for conveying information about vegetation and its environment [van der Maarel 2013: 4].

Vegetation types and their broad-scale distribution, Box & Fujiwara 2013

A vegetation type, or plant community, is identifiable by its distinct appearance compared to other landscape types within a landscape. For example, a grassland and a wetland differ in appearance from each other and from a forest, while a wetland-forest is yet another visibly different vegetation type. Plant species are recognizable by their form, which is related to how the plant functions. For example, in dry environments, plant leaves are more compact with harder surfaces to limit water loss, while plants in wetter environments have larger, “softer” leaves that release water readily when pores open to take in CO2. Such leaves have more surface area for photosynthesis, resulting in faster growth.

This form-function relationship explains why vegetation types differ around the globe. Plant species are adapted to particular climatic conditions according to their proximity to the equator or a coastline, for example, or their elevation.

The geographic regularity of vegetation distribution arises, of course, from the geographic regularity of Earth’s main climatic regions, driven by the global circulation pattern of the Earth’s atmosphere [Box 2013: 466].

Predictive modeling of the potential natural vegetation pattern in northeastern China, Liu et al. 2009

This study uses the concept of Potential Natural Vegetation (PNV), developed in the mid-1900s by German botanist Reinhold Tüxen. Described by the authors as “one of the most successful novelties in vegetation science over the last decades” [Liu 2009: 1313], PNV can be defined as a projection of the natural vegetation that would exist in a given area in the absence of human interference.

“By showing the relationships between environmental variables and vegetation types, maps of the PNV are an important instrument in the study and planning of the environment, and act as decision-support tools for the solutions to environmental issues” [Liu 2009: 1313]. Such maps are informed by studying remnant natural (old growth) vegetation in the area and site observations of the area to be mapped.

Computer modeling can be used to predict “the geographic distribution of vegetation composition across a landscape from mapped environmental variables, such as climate, soils, and geology. When a predictive vegetation modeling is calibrated using observation of vegetation composition taken from mature or ‘climax’ vegetation stands, then potential natural vegetation is portrayed in a predictive map” [Liu 209: 1314].

Focusing on northeastern China, the study identified 16 vegetation types in the region, along with the environmental factors influencing their distribution. Climatic factors included: mean annual temperature, mean temperature of the coldest month, relative humidity, and potential evapotranspiration rate. Topographical factors were elevation and slope.

“Generally, as the elevation increases, the change of temperature and moisture leads to the obvious differentiation phenomenon in vegetation vertical zones. Slope is related to the hydrology (overland and subsurface flow velocity and runoff rate) and potential soil moisture and soil development of a habitat” [Liu 2009: 1315].

They compared the map created by their model to existing vegetation maps of the region. “Visual comparison of the predicted PNV distributions with their actual equivalents indicates a good agreement” [Liu 2009: 1317]. Some modeled vegetation types did not agree with existing maps, however, meaning that “some more important environmental factors may have been missing in the model” [Liu 2009: 1318]. The authors also state that calibrating their model with additional field data on what is currently growing, collected from throughout the region, would improve the model’s accuracy.

The article concludes by stating that a ‘vegetation-environment’ model can help to determine PNV under not only current, but also predicted future environmental conditions.

Conceptualizing communities as natural entities: a philosophical argument with basic and applied implications, Steen et al. 2017

Ecological restoration aims to recreate lost or degraded ecological communities. However, “community” has been a difficult concept to define – should the definition stress dominant species, species interactions, or a subset of strongly interacting species? These authors propose defining community on the basis of co-evolutionary relationships among species.

We propose that an Evolutionary Community is conceptualized as a unique grouping of species, which occur in a given geographic area and are connected by interspecific and abiotic interactions that have evolved over time [Steen 2017: 1021].

By treating communities “as entities that have formed over evolutionary time; this [Evolutionary Community] concept allows for a philosophical platform to help us understand what many conservation and restoration efforts are trying to accomplish” [Steen 2017: 1031]. That is, it offers a way to conceptualize the end goal of a restoration project. A particular evolutionary community could be recreated by assembling the constituent species, resulting in the ecological interactions among the species resuming as before.

What processes cause a group of species to cohere into a community? We argue that the parts of Evolutionary Communities are bound together by interspecific interactions in a shared biotic and abiotic environment, which promote co-evolution and community structure and dynamics. For example, longleaf pine trees are conduits for lightning strikes that ignite a highly flammable understory, often including dropped longleaf pine needles. The resulting ground fires are necessary for reproduction of other species and maintain habitat suitable for others (e.g., gopher tortoises). Gopher tortoises, through the process of burrow creation, provide structure important to other species. The establishment of one or more of the species listed above facilitated the persistence of additional species [Steen 2017: 1025].

Likewise, the demise of one species will negatively affect, or even cause the demise of, other species that depend on it. Thus, the reason to preserve or recreate an integral community is to support the interdependent component species, each of which in turn support the community as a whole.

Bridging ecology and conservation: from ecological networks to ecosystem function, Harvey et al. 2017

This article emphasizes the importance of species interactions as drivers of ecosystem function.

The classic conservation approach is to set aside national parks or to target specific species for protection, based on their rarity or endangered status. However, these approaches can have trade-offs for non-target species, while also potentially failing to protect ecosystem function. The authors, therefore, suggest that species interactions based on their functional significance should be the main focus on conservation efforts.

We propose that a shift in focus from species to interaction networks is necessary to achieve pressing conservation management and restoration ecology goals of conserving biodiversity, ecosystem processes and ultimately landscape-scale delivery of ecosystem services [Harvey 2017: 371].

Species depend on many other species in their communities, either directly or indirectly. An example of indirect dependence is the Phengaris arion butterfly’s need for European rabbits. The butterfly uses ant nests made in the open areas supplied by rabbit grazing for development of its larvae. Thus, no rabbits means no ants, which means no Phengaris arion.

Focusing on species interactions is more meaningful even than measuring species richness (the number of different species), because interactions can disappear – even if both species are present – if either group’s abundance has significantly dropped. The authors offer the example of 59 regionally extinct lepidoptera (butterfly and moth) species of central Europe. Eight of these extinctions were associated with the loss of particular host plant species, which actually occurred after the lepidoptera went extinct.

Focusing on species interactions is more meaningful even than measuring species richness (the number of different species), because interactions can disappear – even if both species are present – if either group’s abundance has significantly dropped.

Thus, strong declines of host plants can have cascading extinction effects on higher trophic levels before the plants actually go extinct, illustrating that interactions can be lost before any actual decline in species richness (plants persisted at low abundance). This illustrates that preserving keystone interactions, rather than species, can be a proactive way to maintain ecosystem integrity in the face of global change instead of allocating resources to already endangered species [Harvey 2017: 372].

There is interdependence among species even between neighboring ecosystems. For example, a manta ray species in the Palmyra Atoll south of Hawaii depends on two species of native trees to maintain its ocean plankton diet. When these trees were replaced with cultivated coconut palms, marine-foraging birds no longer nested on that shore, depriving the coastal waters of the nitrogen runoff from their guano, which had been feeding the plankton population.

The authors recommend that “the main lever to restore or conserve ecological network structure and stability is the management of spatial configuration” [Harvey 2017: 377]. Reflecting on the Palmyra Atoll, for example, it’s clear that a marine conservation plan would be incomplete without considering the nutrient flow from the tree-bird interactions on land.

Interactions among plants and evolution, Thorpe et al. 2011

This review explores the question of whether plant-plant interactions drive evolutionary changes. “If such evolution is common, plant communities are not random assemblages of species.” The topic is under-studied compared to plant interactions with other groups.

Research on plant–consumer, plant–pollinator and plant–disperser interactions has been central to understanding the complex mutualistic and co-dependent interactions among species that structure communities. However, with some notable exceptions, interactions among plants have not been emphasized as processes that contribute to selection and evolution [Thorpe 2011: 730].

“The simplest interactions among plants are direct interactions, such as facilitation, resource competition and allelopathy” [Thorpe 2011: 731]. Facilitation is when one plant protects an adjacent plant, such as from drought and heat by providing shade, for example, or from browsing by being thorny or toxic to herbivores and surrounding the facilitated plant. Allelopathy refers to plants’ release of toxic substances that suppress the growth of another organism, including other plants. In natural communities, any given plant may be interacting with several different plants at the same time.

In natural communities, any given plant may be interacting with several different plants at the same time.

Competition for sunlight, water, and nutrients drives niche differentiation, or the carving out by species of particular spaces or timing within an ecosystem to obtain a share of limited resources. “The exceptionally rich body of ecological literature on the niche is based in part on the idea that competition can drive the evolution of niche differentiation, thus allowing species to coexist” [Thorpe 2011: 732].

Thorpe et al. refer to an example from a 1976 article by Parrish & Bazzaz , who “found that resource partitioning, as estimated from spatial overlap among root systems, was higher in stable prairie communities with a long community history than in early successional old-field communities composed of species without a common history” [Thorpe 2011: 731]. In other words, plants with a long coexistence history more efficiently divvy up resources than do species lacking a common community history.

The primary hypothesis for positive diversity–ecosystem function relationships has been niche ‘complementarity’, the idea that different species or functional groups occupy niches different enough from each other to more fully utilize resources or space, increasing and stabilizing productivity, and making it more difficult for other species to enter the community [Thorpe 2011: 733].

The authors are somewhat inconclusive, however, about what drives niche complementarity (resource partitioning).

We do not yet know whether complementarity is produced by interactions causing evolutionary shifts in niche space (and thus coexistence and more complete resource use) or by sorting of the existing species pool [Thorpe 2011: 733].

Plants can also adapt to one another’s allelopathic substances over time, a fact that contributes to the argument that plant-plant interactions produce evolutionary changes. “Recent experiments raise the possibility that some invaders may exude allelochemicals that are relatively ineffective against neighbors in natural communities, but highly inhibitory to plants in invaded communities” [Thorpe 2011: 734].

Non-native plants reduce abundance, richness, and host specialization in lepidopteran communities, Burghardt et al. 2010

This research evaluates the impact of the invasion of non-native plants in the distribution of lepidopteran (butterfly, skipper, and moth) communities. The authors assert that although the introduction of non-native plants has not resulted in a “global extinction”, they have had a considerable impact on how ecosystems function—they often result in significant bottom-up reductions of energy available in local food webs.

The experiment established four gardens near mature woodlots containing most, if not all, of the native species planted within the treatment. The richness and abundance were then compared for lepidopteran communities found on native versus corresponding non-native congener^[4] species of 13 woody plant genera. For example, the genus Acer (maple) was selected for this study because the native and non-native maples were widespread in that area. In separate plots, the researchers also compared native plants and unrelated (non-congeneric) non-native plants for lepidopteran richness and abundance.

The study found that lepidopterans suffer from the replacement of native plants by non-natives, especially when those non-natives are unrelated to any native plant species. The authors explain that “insect herbivores adapted to the chemical challenges [toxic plant defenses] of particular native hosts may be able to adopt a novel plant species as a host if its phytochemistry is sufficiently similar to the original hosts” [Burghardt 2010: 10]. Over the two-year study, lepidopteran abundance and richness were depressed both on congener and (unrelated) non-congener non-native plants, but especially on the latter.

The study found that lepidopterans suffer from the replacement of native plants by non-natives, especially when those non-natives are unrelated to any native plant species.

Specialist lepidopteran species, which require specific diet and habitat conditions to survive, fared worse on non-natives than did generalists, which can eat a variety of foods and survive in many different habitats. The authors note, for example, that “geographically novel congeners were acceptable hosts to less than half of the generalists and only one fourth of the specialists that we found on native congeners in 2009” [Burghardt 2010: 11]. Only 7% of specialist species used non-congener non-natives as hosts.

The authors argue that the loss of lepidopteran diversity and abundance due to the displacement of native plant species with non-natives can ripple up the food chain, reducing diversity at higher trophic levels. Reduced diversity leads to lower ecosystem productivity and stability, thus disrupting the whole system.

The authors argue that the loss of lepidopteran diversity and abundance due to the displacement of native plant species with non-natives can ripple up the food chain, reducing diversity at higher trophic levels. Reduced diversity leads to lower ecosystem productivity and stability, thus disrupting the whole system.

Because insect herbivores are near the hub of most terrestrial food webs, comprising essential food stuffs for an incredible diversity of insect predators and parasitoids, spiders, amphibians, lizards, rodents, bats, birds, and even higher predators such as foxes and bears, it is particularly important to understand changes wrought by non-native plants on this critical taxon [Burghardt 2010: 13].

Impact of Native Plants on Bird and Butterfly Biodiversity in Suburban Landscapes, Burghardt, Tallamy & Shriver 2008

In this study, the insect and bird populations of six pairs of suburban yards were measured. Each pair contained one conventionally landscaped yard containing native canopy trees and a mixture of native and non-native shrubs, grasses and understory trees; and one yard with native species only (canopy, understory, shrub and grasses). The level of plant diversity was comparable between each of the pair; only the proportion of native species differed. The authors found that:

Avian abundance, diversity, richness, and biomass (particularly bird species of conservation concern) were all greater on native properties. Native nesting birds that are mostly dependent on insect populations to feed their young were more abundant on native properties. Lepidoptera [butterfly and moth species] abundance and diversity were also higher on native properties, suggesting that food availability might account for the differences detected in the bird communities between native and conventionally landscaped sites [Burghardt 2008: 223].

These results support the authors’ hypothesis based on an understanding of the co-evolutionary roots of species interactions.

Theory backed by decades of empirical evidence predicts that up to 90% of all species of insect herbivores can successfully reproduce only on plant lineages with which they have shared an evolutionary history [Burghardt 2008: 220].

Native plants improve breeding and foraging habitat for an insectivorous bird, Narango, Tallamy & Marra 2017

This study examined whether non-native plants in residential Washington DC limited the presence of the Carolina chickadee, a local breeding insectivore.

We predicted that areas with more native plants would support more chickadees, and chickadees would forage more often in the most insect-producing native plants [Narango 2017: 43].

The authors had also considered the possibility that non-native plants could promote increases in other food items (e.g. non-native arthropods), keeping overall prey biomass similar between native and non-native plants. What they found, though, affirmed their prediction: native plants produce more caterpillars, which in turn support more chickadees. In fact, the birds avoided foraging in non-native plants, including non-native species of the same tree genera: the chickadees preferred maples native to the eastern US compared to European-origin maples.

Native plants produce more caterpillars, which in turn support more chickadees.

Native plants were more likely to host a higher biomass of caterpillars compared to non-native plants, and chickadees strongly preferred to forage in native plants that supported the most caterpillars. In addition, chickadees were less likely to breed in yards as the dominance of non-native plants increased [Narango 2017: 42].

Also unique to our study is that we measured the probability of caterpillar occurrence between congeneric species (e.g. native vs. non-native Acer [maple]). This is particularly important considering the popularity and invasive qualities of congeneric species in this region such as Acer platanoides and Quercus acutissima. Although non-native congeners support more caterpillars in comparison to plants unrelated to any native species, congeners had a 47% (CI: 34%–59%) lower probability of having caterpillars compared to native species [Narango 2017: 47].

The authors state that local insects are adapted to local plants, presumably due to their shared co-evolutionary history.

This occurs in part because herbivorous insects have adapted to circumvent the phytochemical defenses of particular plant lineages, resulting in a radiation of specialized plant-insect associations. During urban conversion, native plants are replaced by non-native species with novel chemical, physical, and phenological features for which native herbivorous arthropods have few physiological or behavioral adaptations [Narango 2017: 42].

Do non-native plants contribute to insect declines? Tallamy, Narango & Mitchell 2020

The widespread distribution of plants outside of their native range due to human activity is a significant yet underrecognized cause of global insect decline, according to this article. To illuminate the issue, the authors: “examine the evidence for and against the hypothesis that long term changes in the species composition of plant assemblages have contributed to local and global declines in the abundance and diversity of the insect communities dependent upon those assemblages” [Tallamy 2020: 2].

To be sure, insect conservationists have long noted the importance of habitat containing appropriate native host plants, but the widespread replacement of native host plants with non-native species has yet to penetrate the growing literature on insect declines in any meaningful way [Tallamy 2020: 1].

It is not simply the absence of native plants harms plant-eating insects, however, but also the presence of non-natives. While some insects feed successfully on non-native plants, this is the minority. Most either avoid non-native plants, or do use them and are killed or malnourished by doing so. For example,

Swallowworts (Vincetoxicum spp.) are confamilials of milkweeds (Asclepias spp.) and have become invasive in parts of the northeastern United States. Similar phytochemistry between swallowworts and milkweeds can lead monarch butterflies (Danaus plexxipus) and milkweed beetles (Chrysochus auratus) to fatally mistake these chemically protected plants as hosts. The degree to which Vincetoxicum act as ecological traps for these taxa is likely to become more pronounced as the plants become dominant and displace milkweeds in the landscape [Tallamy 2020: 3].

Species that share a particular environment over hundreds or thousands of years evolve in relation to one another. For plant-eating insects, adapting to certain plants meant developing “traits to detect and tolerate plant defenses over time” [Tallamy 2020: 2]. Most herbivorous insects adapted to only a particular set of plants, specializing in feeding on those plant hosts.

The diet of most insects is constrained to a single plant family in any one habitat or location, with dietary specialization even narrower both in many temperate lineages and hyper-diverse tropical lineages. In fact, diet specialization increases with decreasing latitudes, concurrent with theories of increased plant and animal diversity in the tropics [Tallamy 2020: 2].

When native plants are displaced in the landscape by non-native species, phytophagous [plant-eating] insects typically do not recognize the novel host for feeding or oviposition [egg laying], or may be unable to overcome novel plant defenses. The concurrent loss of native plant hosts and dominance of non-native plants can lead to local extirpation of phytophagous insects and thus to changes in the composition and structure of local food webs [Tallamy 2020: 2].

The most likely successful substitute for a native plant is a non-native plant in the same genus or family.

Non-native congeners [members of the same genus] or confamilials [members of the same family] that are similar in foliar chemistry and nutrition, phenology, and morphology, may occasionally serve as novel hosts for herbivorous insects and support higher diversity and abundance than non-native, non-congeners. However, novel use of congeners may increase larval mortality, extend development or pupation time, reduce biomass, and reduce fitness compared to that of native hosts [Tallamy 2020: 3].

The narrower the native plant diet an insect species has, the less likely to tolerate novel, non-native food sources. However, there are more species of specialist insects than of generalists, meaning a larger proportion of susceptible species. Adaptability to exotic host plants also depends on an insects’ feeding habits.

Insects with chewing (mandibulate) mouthparts are typically more susceptible to defensive secondary metabolites contained in leaf vacuoles than are insects with sucking (haustelate) mouthparts that tap into poorly defended xylem or phloem fluids. Thus, sucking insects find novel non-native plants to be acceptable hosts more often than do chewing species [Tallamy 2020: 4].

Considering that there are more than 4.5 times as many mandibulate insect herbivores as haustelate species, there is reason for concern when non-native plants replace native hosts; the largest guild of insect herbivores is also the most vulnerable to non-native plants and the most valuable to insectivores [Tallamy 2020: 5].

“The dispersal and spread of invasive plants has been driven by global trade networks and colonialism” [Tallamy 2020: 6] and, more specifically, from agroforestry, forestry, agriculture, and horticulture.

Although plants have always distributed themselves around the globe, the increased temporal and spatial mobility of humans has resulted in an extraordinary increase in the rate of plant movements and most species’ introductions have happened in the last 200 years. Habitat is rapidly being converted from coevolved native ecosystems into novel assemblages of plants and animals, making the conversion of native plant communities into plant assemblages dominated by non-native species one of the most ubiquitous threats to biodiversity today. The introduction of non-native plants has completely transformed the composition of present-day plant communities in both natural and human-dominated ecosystems around the globe and the magnitude of introductions is staggering. An estimated 13,168 plant species (about 3.9% of global vascular flora) have been introduced and naturalized beyond their native ranges as a result of human activity [Tallamy 2020: 6].

Global exchange and accumulation of non-native species, van Kleunen et al. 2015

The ecological, economic, and social damage of human-mediated dispersal of species into new regions, where they possess the ability to naturalize (become self-sustaining their new homeland), is one of the defining features of the Anthropocene Epoch. Globally, human activity has led to the naturalization of nearly 13,168 plant species (equal in size to the native European flora). The results from this research provide a baseline for monitoring global changes in biodiversity while highlighting the immediate action that has to be taken to comprehend and determine the spread of alien species on an international scale.

The ecological, economic, and social damage of human-mediated dispersal of species into new regions, where they possess the ability to naturalize (become self-sustaining their new homeland), is one of the defining features of the Anthropocene Epoch.

At least 3.9% of all currently known vascular plant species have become naturalized outside their natural ranges as a result of human activity. With the continued practice of international traffic and trade and globalization, the likelihood of more and more species being introduced and getting naturalized outside their native range is high.

To assess the accumulation of naturalized species in each continent as well as which continents have been the major donors of alien naturalized plant species globally, the researchers used a novel database, Global Naturalized Alien Flora (GloNAF), in addition to the data on the origin of naturalized species and estimates of the number of native species per continent. They found that when not taking into account the differences in total area, North America has accumulated the highest number of naturalized species (n=5,958). However, when considering the difference in total area, Australasia (a region comprising Australia, New Zealand, and neighboring islands) was found to have more extra-continental species than North America.

One possible explanation is that Australia’s long biogeographical isolation and drying climate have resulted in a native flora that is phylogenetically distinct, but not well-adapted to exploit the novel habitats created by European settlers [van Kleunen 2015: 101].

The major donors of alien species are Europe and temperate Asia, while North America is also a significant donor.

Ecological and evolutionary consequences of biotic homogenization, Olden et al. 2004

Anthropogenic environmental change and global dispersal of a wide variety of species outside their native ranges has expanded the range of “cosmopolitan,” non-native species and shrunk the range of regional and endemic species. “This replacement of specific native forms by generalist non-natives in space and time has mixed the taxonomic composition of once disparate biotas, an occurrence termed ‘biotic homogenization’” [Olden 2004: 18].

The authors explore the effect of this “global erosion of regional distinctiveness” [Olden 2004: 18] at three levels: Genetic homogenization reduces genetic variability within species or among populations of species, while taxonomic homogenization reduces distinctiveness among communities. Functional homogenization refers to a reduction of functional traits within an ecosystem. The identity of species making up a community, along with their respective functional traits, determines “ecosystem functions (such as nutrient retention or energy flow)” [Olden 2004: 20], so that narrowing species compositions risks diminishing ecosystem function.

A decrease in functional diversity might reduce overall community and ecosystem functioning, stability and resistance to environmental change by simply narrowing the available range of species-specific responses. Consider a severe drought that strongly affects a subset of species in a community that has (or lacks) a particular suite of functional traits. Historical communities, with much greater breadth in functional space, should exhibit higher resistance or resilience when compared with homogenized communities [Olden 2004: 20].

Genetic homogenization occurs when two distinct locally adapted populations of the same species interbreed. It also occurs when a single variety (such as captive fish bred in a central location) are released in many places to replenish dwindling native stocks. While such mixing has the potential to increase species diversity, this outcome is not assured.

Intraspecific hybridization can homogenize the unique characteristics of geographically distinct populations, as well as compromise the fitness of individuals by disrupting local adaptations [Olden 2004: 19].

Linking Restoration and Ecological Succession, Walker, Walker & Hobbs (eds) 2007

This book draws lessons from ecological succession theory to inform ecological restoration, stating that: “restoration is fundamentally the management of succession” [Walker 2007: vi]. The latter is the natural process by which plants first colonize “new” land (post landslide, glacial retreat or volcanic eruption, for example) or degraded land, and over time develop into mature ecosystems through a series of changing plant communities. Ecological restoration is a human-led initiative to restore functioning ecosystems, or at least vegetation, on land degraded through human activity. The ultimate goal of restoration is to “establish a self-sufficient ecosystem that requires minimal or no continuing human inputs in order to provide a continuing supply of goods and services” [Hobbs 2007: 177].

Effective ecosystem restoration requires ecological knowledge. Likewise, the outcomes of such projects demonstrate our comprehension, or lack thereof, of ecological concepts: “Restoration is the acid test of our ability to understand not only how ecosystems are assembled and held together but also how they change over time” [Walker 2007: vi]. The authors contend, however, that restoration projects are more often guided by engineering, horticulture, and agronomy than by ecology. Aiming to clarify the ways in which ecological succession theory can and should inform restoration, this book poses the question: “What is the minimum amount of biophysical and successional information needed to restore a specific landscape or area” [Walker 2007b: 2]?

Succession comprises many ecological processes that underpin all ecological restoration and ecological restoration is a manipulation of these processes to achieve its goals. This means it is essential to understand how succession operates, and when and how to manipulate it [Prach 2007: 121].

Restoration can explicitly embrace a hands-off approach, where land is simply left to repair itself through natural ecological succession. On the other hand, understanding the successional process allows manipulation of various stages to speed up the process. For example, in the first stage of primary succession “winds deposit dust, pollen, seeds, and insects crucial to reducing infertility” [del Moral 2007: 23], on bare, inhospitable ground. Tough pioneer plants are able to establish then create shade, trap sediment, and deposit organic matter when they die, creating slightly better conditions for the next wave of colonizing plants. To mimic this first stage of site “amelioration”, the site can be physically manipulated by reshaping the ground for improved drainage or adding organic matter, for example.

Biological manipulation involves sowing or planting local/native varieties of later successional species that may not be otherwise present in the area due to human transformation of the broader landscape. While earlier successional species tend to have small, easily transported seeds, the later successional species (such as large canopy trees) that are often the target of restoration efforts often have large, less mobile seeds. Thus, if those plants are not present in the immediate environment as seed stock, they may never establish in the restored site without human assistance.

Near-Natural Silviculture: Sustainable Approach for Urban Re-naturalization Assessment Based on 10 Years Recovering Dynamics and Eco-Benefits in Shanghai, Guo et. al 2015

As one of China’s major cities, Shanghai’s natural sub-ecosystem^[5] has suffered drastic damage due to human activities and urbanization. Although urban re-naturalization has gained attention from city leaders, urban tree planting has largely consisted of two methods with limited ecological potential. One favors fast-growing monocultures to produce timber products and other benefits, while the other approach is to plant non-native species for decorative purposes. The authors believe the restoration progress of the natural sub-ecosystem could be further improved by adopting the “near-natural” method based on the concepts of potential natural vegetation^[6] and ecological succession.

The near-natural forest uses all native species and aims to create a complex structure with high biodiversity, high biomass and multiplayer canopies. It was adopted successfully in many countries, but the authors thought long-term studies of these forests were lacking. Therefore, they conducted a 10-year study at a near-natural forest established in 2000 in Pudong New Area of Shanghai to investigate the effectiveness of the forest in providing ecological benefits.

Results showed that the near-natural forest had higher sustainability value than artificial (“even-aged, managed”) forest in Shanghai based on its ecological and economic benefits. The high tree density and multiple vertical structures of the forest improved the air quality and soil fertility and decreased the concentrations of air bacteria and dust. It also had a much lower planting and maintenance cost than artificial traditional methods. Although the near-natural forest could not transcend the benefits of the natural forest, the study successfully proved its important role in urban re-naturalization by bridging the difference between the artificial and natural forests.

Results showed that the near-natural forest had higher sustainability value than artificial (“even-aged, managed”) forest in Shanghai based on its ecological and economic benefits

Over the course of the study, authors discovered a potential limitation of the approach, at least in its application in Shanghai. They observed high evergreen seedlings mortality, attributable to over-exposure to sunlight. Therefore, in subsequent plantings in 2003 and 2004, the authors modified the approach to optimize it to local conditions:

The key to the new method is to create a mixed deciduous–evergreen community by simultaneously planting shade-tolerant evergreen broad-leaved species and light-demanding deciduous broad-leaved species, but using smaller individuals for the former and larger individuals for the latter to form a multilayer vegetation structure. The shade-tolerant evergreen species benefit from the rapid growth of the light-demanding deciduous species, which offer shade and nutrients in the form of litter layer-based fertilizer, improving the soil for the evergreen species [Guo 2015: 5].

Overall, they suggest that the near-natural forest is a very sustainable method to be applied in Shanghai.

Tree planting is not a simple science, Holl & Brancalion 2020

Well-planned tree-planting projects are an important component of global efforts to improve ecological and human well-being. But tree planting becomes problematic when it is promoted as a simple, silver bullet solution and overshadows other actions that have greater potential for addressing the drivers of specific environmental problems, such as taking bold and rapid steps to reduce deforestation and greenhouse gas emissions [Holl 2020: 580].

Some of the pitfalls to avoid in tree planting initiatives, according to the authors, include:

• Use of non-native species, which does not result in a true forest and can result in ground water depletion in arid environments.
• Planting trees in historic grasslands and savannas, harming those native ecosystems and species.
• Abandoning trees after they are planted, which can result in high mortality due to insufficient water for developing saplings, being shaded out by faster growing herbaceous plants, grazing, or being re-cleared.
• Planting trees in agricultural land, which risks pushing crop production into native forest land, which is then deforested.

The authors insist that reforestation takes careful planning, stakeholder engagement, clear goal-setting, and long-term monitoring and adaptive management of planted tree stands to ensure their survival. Above all, existing mature, native forests should be preserved.

The authors insist that reforestation takes careful planning, stakeholder engagement, clear goal-setting, and long-term monitoring and adaptive management of planted tree stands to ensure their survival. Above all, existing mature, native forests should be preserved.

The first priority to increase the overall number of trees on the planet must be to reduce the current rapid rate of forest clearing and degradation in many areas of the world. The immediate response of the G7 nations to the 2019 Amazon fires was to offer funding to reforest these areas, rather than to address the core issues of enforcing laws, protecting lands of indigenous people, and providing incentives to landowners to maintain forest cover. The simplistic assumption that tree planting can immediately compensate for clearing intact forest is not uncommon. Nonetheless, a large body of literature shows that even the best-planned restoration projects rarely fully recover the biodiversity of intact forest, owing to a lack of sources of forest-dependent flora and fauna in deforested landscapes, as well as degraded abiotic conditions resulting from anthropogenic activities [Holl 2020: 581].

Produced by the Native Plant Society of Oregon, this article argues that, while the use of native species is an accepted tenet of conservation, the term “native” is not necessarily well understood; they attempt to clarify the term.

“Any definition of a native species, native ecosystem, or native landscape requires an historical benchmark” [Wilson 1991: 16]. Over the past 20,000 years, “vegetation in the Willamette Valley has changed dramatically with changing climate. Vegetation in a single place has probably varied from boreal parkland, to conifer forest, to oak savanna, to prairie. Each climatic phase supported a different flora” [Wilson 1991: 16]. Each of these vegetation types was native to a particular place, according to particular climatic conditions that changed overtime. The vegetation that developed in the past 10,000 years – the current Holocene period of climate stability – is thus the relevant reference.

“For the Pacific Northwest, the period that ended with Euro-American settlement is a natural historical benchmark. This period lasted long enough to have a significant impact on the vegetation of the region. The climates of much earlier times were different enough to limit their usefulness in defining today’s ecosystems” [Wilson 1991: 16]. Thus, “any species that had occurred in a particular ecological habitat [of the Pacific Northwest] before Euro-American settlement is a species native to that habitat” [Wilson 1991: 17].

A native ecosystem, then, is one dominated by native plants, animals and microorganisms that occurred together before the time of Euro-American settlement. Key species – for example, the dominant photosynthesizing plants, the top carnivores, the important decomposers, the nitrogen-fixers – must be present for a native ecosystem to persist and function on its own. To artificially maintain a conserved or restored ecosystem without all of its crucial components is both difficult and expensive. The species of native ecosystem must also occur together in nature. For example, landscaping with an artificial mixture of native species like vine maple, blue bunch wheatgrass, and Jeffrey pine does not produce a native ecosystem. These species are native to different areas within Oregon, but they would not naturally grow together in the same ecosystem. Restoration of native ecosystems must also account for proper structure and appearance. For example, a red fescue lawn does not have the structural complexity and species diversity exhibited by native bunchgrass prairies [Wilson 1991: 17].

Key species – for example, the dominant photosynthesizing plants, the top carnivores, the important decomposers, the nitrogen-fixers – must be present for a native ecosystem to persist and function on its own. To artificially maintain a conserved or restored ecosystem without all of its crucial components is both difficult and expensive [Wilson 1991: 17].

These authors define the concept of a plant community through discussion of its evolution. They start by defining the term ‘vegetation’ in a way that may surprise some readers because it excludes plants growing in certain situations. To be considered vegetation, plants need to emerge spontaneously.

Vegetation, the central object of study in vegetation ecology, can be loosely defined as a system of largely spontaneously growing plants. Not all growing plants form vegetation, for instance, a sown corn field or a flower bed in a garden do not. But the weeds surrounding such plants do form vegetation. A pine plantation will become vegetation after some years of spontaneous growth of the pine trees and the subsequent development of an understory [van der Maarel 2013: 1].

Two competing schools of thought regarding the nature of a stand of plants growing together geographically are represented by two early 20th Century botanists. H.A. Gleason observed “that species are ‘individualistically’ distributed along omnipresent environmental gradients and thus cannot form bounded communities” [van der Maarel 2013: 2]. By contrast, E. Clements compared plant community with an integral organism, where the whole was greater than the sum of its parts. During the same time period, the Braun-Blanquet approach was developed, which “paid much attention to the relations of plant communities with the environment and the interactions within communities, which is now incorporated in the concept of ecosystem” [van der Maarel 2013: 2].

The authors state that while individual plant species are distributed according to abiotic environmental conditions, the fact of being co-located with particular sets of other species in a particular environment results in interspecies interactions, which are in fact ecosystem processes (emergent properties).

In conclusion, a plant community is generally recognized as a relatively uniform piece of vegetation in a uniform environment, with a recognizable floristic composition and structure, that is relatively distinct from the surrounding vegetation. Even if the populations of the participating species are usually distributed individualistically in the landscape, they may well interact within the community and build up an integrated unit with emergent properties. At the same time, plant communities can be convenient units for conveying information about vegetation and its environment [van der Maarel 2013: 4].

A vegetation type, or plant community, is identifiable by its distinct appearance compared to other landscape types within a landscape. For example, a grassland and a wetland differ in appearance from each other and from a forest, while a wetland-forest is yet another visibly different vegetation type. Plant species are recognizable by their form, which is related to how the plant functions. For example, in dry environments, plant leaves are more compact with harder surfaces to limit water loss, while plants in wetter environments have larger, “softer” leaves that release water readily when pores open to take in CO2. Such leaves have more surface area for photosynthesis, resulting in faster growth.

This form-function relationship explains why vegetation types differ around the globe. Plant species are adapted to particular climatic conditions according to their proximity to the equator or a coastline, for example, or their elevation.

The geographic regularity of vegetation distribution arises, of course, from the geographic regularity of Earth’s main climatic regions, driven by the global circulation pattern of the Earth’s atmosphere [Box 2013: 466].

This study uses the concept of Potential Natural Vegetation (PNV), developed in the mid-1900s by German botanist Reinhold Tüxen. Described by the authors as “one of the most successful novelties in vegetation science over the last decades” [Liu 2009: 1313], PNV can be defined as a projection of the natural vegetation that would exist in a given area in the absence of human interference.

“By showing the relationships between environmental variables and vegetation types, maps of the PNV are an important instrument in the study and planning of the environment, and act as decision-support tools for the solutions to environmental issues” [Liu 2009: 1313]. Such maps are informed by studying remnant natural (old growth) vegetation in the area and site observations of the area to be mapped.

Computer modeling can be used to predict “the geographic distribution of vegetation composition across a landscape from mapped environmental variables, such as climate, soils, and geology. When a predictive vegetation modeling is calibrated using observation of vegetation composition taken from mature or ‘climax’ vegetation stands, then potential natural vegetation is portrayed in a predictive map” [Liu 209: 1314].

Focusing on northeastern China, the study identified 16 vegetation types in the region, along with the environmental factors influencing their distribution. Climatic factors included: mean annual temperature, mean temperature of the coldest month, relative humidity, and potential evapotranspiration rate. Topographical factors were elevation and slope.

“Generally, as the elevation increases, the change of temperature and moisture leads to the obvious differentiation phenomenon in vegetation vertical zones. Slope is related to the hydrology (overland and subsurface flow velocity and runoff rate) and potential soil moisture and soil development of a habitat” [Liu 2009: 1315].

They compared the map created by their model to existing vegetation maps of the region. “Visual comparison of the predicted PNV distributions with their actual equivalents indicates a good agreement” [Liu 2009: 1317]. Some modeled vegetation types did not agree with existing maps, however, meaning that “some more important environmental factors may have been missing in the model” [Liu 2009: 1318]. The authors also state that calibrating their model with additional field data on what is currently growing, collected from throughout the region, would improve the model’s accuracy.

The article concludes by stating that a ‘vegetation-environment’ model can help to determine PNV under not only current, but also predicted future environmental conditions.

This review explores the question of whether plant-plant interactions drive evolutionary changes. “If such evolution is common, plant communities are not random assemblages of species.” The topic is under-studied compared to plant interactions with other groups.

Research on plant–consumer, plant–pollinator and plant–disperser interactions has been central to understanding the complex mutualistic and co-dependent interactions among species that structure communities. However, with some notable exceptions, interactions among plants have not been emphasized as processes that contribute to selection and evolution [Thorpe 2011: 730].

“The simplest interactions among plants are direct interactions, such as facilitation, resource competition and allelopathy” [Thorpe 2011: 731]. Facilitation is when one plant protects an adjacent plant, such as from drought and heat by providing shade, for example, or from browsing by being thorny or toxic to herbivores and surrounding the facilitated plant. Allelopathy refers to plants’ release of toxic substances that suppress the growth of another organism, including other plants. In natural communities, any given plant may be interacting with several different plants at the same time.

In natural communities, any given plant may be interacting with several different plants at the same time.

Competition for sunlight, water, and nutrients drives niche differentiation, or the carving out by species of particular spaces or timing within an ecosystem to obtain a share of limited resources. “The exceptionally rich body of ecological literature on the niche is based in part on the idea that competition can drive the evolution of niche differentiation, thus allowing species to coexist” [Thorpe 2011: 732].

Thorpe et al. refer to an example from a 1976 article by Parrish & Bazzaz , who “found that resource partitioning, as estimated from spatial overlap among root systems, was higher in stable prairie communities with a long community history than in early successional old-field communities composed of species without a common history” [Thorpe 2011: 731]. In other words, plants with a long coexistence history more efficiently divvy up resources than do species lacking a common community history.

The primary hypothesis for positive diversity–ecosystem function relationships has been niche ‘complementarity’, the idea that different species or functional groups occupy niches different enough from each other to more fully utilize resources or space, increasing and stabilizing productivity, and making it more difficult for other species to enter the community [Thorpe 2011: 733].

The authors are somewhat inconclusive, however, about what drives niche complementarity (resource partitioning).

We do not yet know whether complementarity is produced by interactions causing evolutionary shifts in niche space (and thus coexistence and more complete resource use) or by sorting of the existing species pool [Thorpe 2011: 733].

Plants can also adapt to one another’s allelopathic substances over time, a fact that contributes to the argument that plant-plant interactions produce evolutionary changes. “Recent experiments raise the possibility that some invaders may exude allelochemicals that are relatively ineffective against neighbors in natural communities, but highly inhibitory to plants in invaded communities” [Thorpe 2011: 734].

In this study, the insect and bird populations of six pairs of suburban yards were measured. Each pair contained one conventionally landscaped yard containing native canopy trees and a mixture of native and non-native shrubs, grasses and understory trees; and one yard with native species only (canopy, understory, shrub and grasses). The level of plant diversity was comparable between each of the pair; only the proportion of native species differed. The authors found that:

Avian abundance, diversity, richness, and biomass (particularly bird species of conservation concern) were all greater on native properties. Native nesting birds that are mostly dependent on insect populations to feed their young were more abundant on native properties. Lepidoptera [butterfly and moth species] abundance and diversity were also higher on native properties, suggesting that food availability might account for the differences detected in the bird communities between native and conventionally landscaped sites [Burghardt 2008: 223].

These results support the authors’ hypothesis based on an understanding of the co-evolutionary roots of species interactions.

Theory backed by decades of empirical evidence predicts that up to 90% of all species of insect herbivores can successfully reproduce only on plant lineages with which they have shared an evolutionary history [Burghardt 2008: 220].

This study examined whether non-native plants in residential Washington DC limited the presence of the Carolina chickadee, a local breeding insectivore.

We predicted that areas with more native plants would support more chickadees, and chickadees would forage more often in the most insect-producing native plants [Narango 2017: 43].

The authors had also considered the possibility that non-native plants could promote increases in other food items (e.g. non-native arthropods), keeping overall prey biomass similar between native and non-native plants. What they found, though, affirmed their prediction: native plants produce more caterpillars, which in turn support more chickadees. In fact, the birds avoided foraging in non-native plants, including non-native species of the same tree genera: the chickadees preferred maples native to the eastern US compared to European-origin maples.

Native plants produce more caterpillars, which in turn support more chickadees.

Native plants were more likely to host a higher biomass of caterpillars compared to non-native plants, and chickadees strongly preferred to forage in native plants that supported the most caterpillars. In addition, chickadees were less likely to breed in yards as the dominance of non-native plants increased [Narango 2017: 42].

Also unique to our study is that we measured the probability of caterpillar occurrence between congeneric species (e.g. native vs. non-native Acer [maple]). This is particularly important considering the popularity and invasive qualities of congeneric species in this region such as Acer platanoides and Quercus acutissima. Although non-native congeners support more caterpillars in comparison to plants unrelated to any native species, congeners had a 47% (CI: 34%–59%) lower probability of having caterpillars compared to native species [Narango 2017: 47].

The authors state that local insects are adapted to local plants, presumably due to their shared co-evolutionary history.

This occurs in part because herbivorous insects have adapted to circumvent the phytochemical defenses of particular plant lineages, resulting in a radiation of specialized plant-insect associations. During urban conversion, native plants are replaced by non-native species with novel chemical, physical, and phenological features for which native herbivorous arthropods have few physiological or behavioral adaptations [Narango 2017: 42].

The widespread distribution of plants outside of their native range due to human activity is a significant yet underrecognized cause of global insect decline, according to this article. To illuminate the issue, the authors: “examine the evidence for and against the hypothesis that long term changes in the species composition of plant assemblages have contributed to local and global declines in the abundance and diversity of the insect communities dependent upon those assemblages” [Tallamy 2020: 2].

To be sure, insect conservationists have long noted the importance of habitat containing appropriate native host plants, but the widespread replacement of native host plants with non-native species has yet to penetrate the growing literature on insect declines in any meaningful way [Tallamy 2020: 1].

It is not simply the absence of native plants harms plant-eating insects, however, but also the presence of non-natives. While some insects feed successfully on non-native plants, this is the minority. Most either avoid non-native plants, or do use them and are killed or malnourished by doing so. For example,

Swallowworts (Vincetoxicum spp.) are confamilials of milkweeds (Asclepias spp.) and have become invasive in parts of the northeastern United States. Similar phytochemistry between swallowworts and milkweeds can lead monarch butterflies (Danaus plexxipus) and milkweed beetles (Chrysochus auratus) to fatally mistake these chemically protected plants as hosts. The degree to which Vincetoxicum act as ecological traps for these taxa is likely to become more pronounced as the plants become dominant and displace milkweeds in the landscape [Tallamy 2020: 3].

Species that share a particular environment over hundreds or thousands of years evolve in relation to one another. For plant-eating insects, adapting to certain plants meant developing “traits to detect and tolerate plant defenses over time” [Tallamy 2020: 2]. Most herbivorous insects adapted to only a particular set of plants, specializing in feeding on those plant hosts.

The diet of most insects is constrained to a single plant family in any one habitat or location, with dietary specialization even narrower both in many temperate lineages and hyper-diverse tropical lineages. In fact, diet specialization increases with decreasing latitudes, concurrent with theories of increased plant and animal diversity in the tropics [Tallamy 2020: 2].

When native plants are displaced in the landscape by non-native species, phytophagous [plant-eating] insects typically do not recognize the novel host for feeding or oviposition [egg laying], or may be unable to overcome novel plant defenses. The concurrent loss of native plant hosts and dominance of non-native plants can lead to local extirpation of phytophagous insects and thus to changes in the composition and structure of local food webs [Tallamy 2020: 2].

The most likely successful substitute for a native plant is a non-native plant in the same genus or family.

Non-native congeners [members of the same genus] or confamilials [members of the same family] that are similar in foliar chemistry and nutrition, phenology, and morphology, may occasionally serve as novel hosts for herbivorous insects and support higher diversity and abundance than non-native, non-congeners. However, novel use of congeners may increase larval mortality, extend development or pupation time, reduce biomass, and reduce fitness compared to that of native hosts [Tallamy 2020: 3].

The narrower the native plant diet an insect species has, the less likely to tolerate novel, non-native food sources. However, there are more species of specialist insects than of generalists, meaning a larger proportion of susceptible species. Adaptability to exotic host plants also depends on an insects’ feeding habits.

Insects with chewing (mandibulate) mouthparts are typically more susceptible to defensive secondary metabolites contained in leaf vacuoles than are insects with sucking (haustelate) mouthparts that tap into poorly defended xylem or phloem fluids. Thus, sucking insects find novel non-native plants to be acceptable hosts more often than do chewing species [Tallamy 2020: 4].

Considering that there are more than 4.5 times as many mandibulate insect herbivores as haustelate species, there is reason for concern when non-native plants replace native hosts; the largest guild of insect herbivores is also the most vulnerable to non-native plants and the most valuable to insectivores [Tallamy 2020: 5].

“The dispersal and spread of invasive plants has been driven by global trade networks and colonialism” [Tallamy 2020: 6] and, more specifically, from agroforestry, forestry, agriculture, and horticulture.

Although plants have always distributed themselves around the globe, the increased temporal and spatial mobility of humans has resulted in an extraordinary increase in the rate of plant movements and most species’ introductions have happened in the last 200 years. Habitat is rapidly being converted from coevolved native ecosystems into novel assemblages of plants and animals, making the conversion of native plant communities into plant assemblages dominated by non-native species one of the most ubiquitous threats to biodiversity today. The introduction of non-native plants has completely transformed the composition of present-day plant communities in both natural and human-dominated ecosystems around the globe and the magnitude of introductions is staggering. An estimated 13,168 plant species (about 3.9% of global vascular flora) have been introduced and naturalized beyond their native ranges as a result of human activity [Tallamy 2020: 6].

The ecological, economic, and social damage of human-mediated dispersal of species into new regions, where they possess the ability to naturalize (become self-sustaining their new homeland), is one of the defining features of the Anthropocene Epoch. Globally, human activity has led to the naturalization of nearly 13,168 plant species (equal in size to the native European flora). The results from this research provide a baseline for monitoring global changes in biodiversity while highlighting the immediate action that has to be taken to comprehend and determine the spread of alien species on an international scale.

The ecological, economic, and social damage of human-mediated dispersal of species into new regions, where they possess the ability to naturalize (become self-sustaining their new homeland), is one of the defining features of the Anthropocene Epoch.

At least 3.9% of all currently known vascular plant species have become naturalized outside their natural ranges as a result of human activity. With the continued practice of international traffic and trade and globalization, the likelihood of more and more species being introduced and getting naturalized outside their native range is high.

To assess the accumulation of naturalized species in each continent as well as which continents have been the major donors of alien naturalized plant species globally, the researchers used a novel database, Global Naturalized Alien Flora (GloNAF), in addition to the data on the origin of naturalized species and estimates of the number of native species per continent. They found that when not taking into account the differences in total area, North America has accumulated the highest number of naturalized species (n=5,958). However, when considering the difference in total area, Australasia (a region comprising Australia, New Zealand, and neighboring islands) was found to have more extra-continental species than North America.

One possible explanation is that Australia’s long biogeographical isolation and drying climate have resulted in a native flora that is phylogenetically distinct, but not well-adapted to exploit the novel habitats created by European settlers [van Kleunen 2015: 101].

The major donors of alien species are Europe and temperate Asia, while North America is also a significant donor.

This book draws lessons from ecological succession theory to inform ecological restoration, stating that: “restoration is fundamentally the management of succession” [Walker 2007: vi]. The latter is the natural process by which plants first colonize “new” land (post landslide, glacial retreat or volcanic eruption, for example) or degraded land, and over time develop into mature ecosystems through a series of changing plant communities. Ecological restoration is a human-led initiative to restore functioning ecosystems, or at least vegetation, on land degraded through human activity. The ultimate goal of restoration is to “establish a self-sufficient ecosystem that requires minimal or no continuing human inputs in order to provide a continuing supply of goods and services” [Hobbs 2007: 177].

Effective ecosystem restoration requires ecological knowledge. Likewise, the outcomes of such projects demonstrate our comprehension, or lack thereof, of ecological concepts: “Restoration is the acid test of our ability to understand not only how ecosystems are assembled and held together but also how they change over time” [Walker 2007: vi]. The authors contend, however, that restoration projects are more often guided by engineering, horticulture, and agronomy than by ecology. Aiming to clarify the ways in which ecological succession theory can and should inform restoration, this book poses the question: “What is the minimum amount of biophysical and successional information needed to restore a specific landscape or area” [Walker 2007b: 2]?

Succession comprises many ecological processes that underpin all ecological restoration and ecological restoration is a manipulation of these processes to achieve its goals. This means it is essential to understand how succession operates, and when and how to manipulate it [Prach 2007: 121].

Restoration can explicitly embrace a hands-off approach, where land is simply left to repair itself through natural ecological succession. On the other hand, understanding the successional process allows manipulation of various stages to speed up the process. For example, in the first stage of primary succession “winds deposit dust, pollen, seeds, and insects crucial to reducing infertility” [del Moral 2007: 23], on bare, inhospitable ground. Tough pioneer plants are able to establish then create shade, trap sediment, and deposit organic matter when they die, creating slightly better conditions for the next wave of colonizing plants. To mimic this first stage of site “amelioration”, the site can be physically manipulated by reshaping the ground for improved drainage or adding organic matter, for example.

Biological manipulation involves sowing or planting local/native varieties of later successional species that may not be otherwise present in the area due to human transformation of the broader landscape. While earlier successional species tend to have small, easily transported seeds, the later successional species (such as large canopy trees) that are often the target of restoration efforts often have large, less mobile seeds. Thus, if those plants are not present in the immediate environment as seed stock, they may never establish in the restored site without human assistance.

As one of China’s major cities, Shanghai’s natural sub-ecosystem^[5] has suffered drastic damage due to human activities and urbanization. Although urban re-naturalization has gained attention from city leaders, urban tree planting has largely consisted of two methods with limited ecological potential. One favors fast-growing monocultures to produce timber products and other benefits, while the other approach is to plant non-native species for decorative purposes. The authors believe the restoration progress of the natural sub-ecosystem could be further improved by adopting the “near-natural” method based on the concepts of potential natural vegetation^[6] and ecological succession.

The near-natural forest uses all native species and aims to create a complex structure with high biodiversity, high biomass and multiplayer canopies. It was adopted successfully in many countries, but the authors thought long-term studies of these forests were lacking. Therefore, they conducted a 10-year study at a near-natural forest established in 2000 in Pudong New Area of Shanghai to investigate the effectiveness of the forest in providing ecological benefits.

Results showed that the near-natural forest had higher sustainability value than artificial (“even-aged, managed”) forest in Shanghai based on its ecological and economic benefits. The high tree density and multiple vertical structures of the forest improved the air quality and soil fertility and decreased the concentrations of air bacteria and dust. It also had a much lower planting and maintenance cost than artificial traditional methods. Although the near-natural forest could not transcend the benefits of the natural forest, the study successfully proved its important role in urban re-naturalization by bridging the difference between the artificial and natural forests.

Results showed that the near-natural forest had higher sustainability value than artificial (“even-aged, managed”) forest in Shanghai based on its ecological and economic benefits

Over the course of the study, authors discovered a potential limitation of the approach, at least in its application in Shanghai. They observed high evergreen seedlings mortality, attributable to over-exposure to sunlight. Therefore, in subsequent plantings in 2003 and 2004, the authors modified the approach to optimize it to local conditions:

The key to the new method is to create a mixed deciduous–evergreen community by simultaneously planting shade-tolerant evergreen broad-leaved species and light-demanding deciduous broad-leaved species, but using smaller individuals for the former and larger individuals for the latter to form a multilayer vegetation structure. The shade-tolerant evergreen species benefit from the rapid growth of the light-demanding deciduous species, which offer shade and nutrients in the form of litter layer-based fertilizer, improving the soil for the evergreen species [Guo 2015: 5].

Overall, they suggest that the near-natural forest is a very sustainable method to be applied in Shanghai.

Well-planned tree-planting projects are an important component of global efforts to improve ecological and human well-being. But tree planting becomes problematic when it is promoted as a simple, silver bullet solution and overshadows other actions that have greater potential for addressing the drivers of specific environmental problems, such as taking bold and rapid steps to reduce deforestation and greenhouse gas emissions [Holl 2020: 580].

Some of the pitfalls to avoid in tree planting initiatives, according to the authors, include:

• Use of non-native species, which does not result in a true forest and can result in ground water depletion in arid environments.
• Planting trees in historic grasslands and savannas, harming those native ecosystems and species.
• Abandoning trees after they are planted, which can result in high mortality due to insufficient water for developing saplings, being shaded out by faster growing herbaceous plants, grazing, or being re-cleared.
• Planting trees in agricultural land, which risks pushing crop production into native forest land, which is then deforested.

The authors insist that reforestation takes careful planning, stakeholder engagement, clear goal-setting, and long-term monitoring and adaptive management of planted tree stands to ensure their survival. Above all, existing mature, native forests should be preserved.

The authors insist that reforestation takes careful planning, stakeholder engagement, clear goal-setting, and long-term monitoring and adaptive management of planted tree stands to ensure their survival. Above all, existing mature, native forests should be preserved.

The first priority to increase the overall number of trees on the planet must be to reduce the current rapid rate of forest clearing and degradation in many areas of the world. The immediate response of the G7 nations to the 2019 Amazon fires was to offer funding to reforest these areas, rather than to address the core issues of enforcing laws, protecting lands of indigenous people, and providing incentives to landowners to maintain forest cover. The simplistic assumption that tree planting can immediately compensate for clearing intact forest is not uncommon. Nonetheless, a large body of literature shows that even the best-planned restoration projects rarely fully recover the biodiversity of intact forest, owing to a lack of sources of forest-dependent flora and fauna in deforested landscapes, as well as degraded abiotic conditions resulting from anthropogenic activities [Holl 2020: 581].

Connectivity conservation pays attention to landscape connectivity to support animal species’ movements, keep ecological processes intact, and promote biodiversity. While the strategy of conserving connected, non-fragmented areas and respecting animals’ movement patterns is sound, in practice these plans are usually designed around a single species and its needs.

Brennan et al. looked at the limitations of a single-species focus, and evaluated the movement patterns of multiple species. They created connectivity maps for six large mammal species in the Kavango-Zambezi (KAZA) transfrontier conservation area straddling Angola, Zambia, Zimbabwe, Botswana, and Namibia, and assessed how each individual species’ connectivity maps correlated with that of the others.

This then allowed the authors to identify good ‘surrogate species for connectivity’ – that is, species whose connectivity maps were good representations of other species’ movements through the same area. They also took a look at different types of barriers to animal movements and determined that fences were the greatest obstacle to movement, while roads, rivers, and human-settled areas also deterred movement. Finally, they identified connectivity hotspots on the landscape, which are like bottlenecks through which multiple species pass due to barriers elsewhere. These connectivity hotspots are thus essential places to focus conservation efforts.

The researchers found the hyena and African wild dog to be the most apt surrogate species for connectivity, in spite of a popular practice of using elephants to determine the geographic targets of conservation efforts.

In our examination of connectivity across the landscape, female elephants were found to be only weakly correlated with the five other species in our study. Spotted hyena and African wild dog, in contrast, were strongly correlated with the greatest number of species. They also appeared to be complementary surrogates (i.e. they were correlated with different species), in which case combining their connectivity models could further extend the relevancy of connectivity conservation plans to other species. Thus, as both species are also charismatic, wide-ranging species of conservation concern, they may represent good umbrella species for connectivity in the KAZA region [Brennan 2020: 1707].

They went on to say that “while elephants may not be good surrogate species for connectivity across entire landscapes, they may still be effective as a surrogate at local scales where they can help protect local movement pathways or stepping-stone habitats for other species” [Brennan 2020: 1707].

Their conclusion is not that we should stop paying attention to elephants, which serve important ecological functions and are an iconic and culturally significant animal. Rather, we should look for gaps that may arise if we only conserve areas based on elephant movements, and put these techniques of comparing and combining different species’ movement patterns to use. Noting that animal movements and ecological dynamics play out at different scales, from entire landscapes and transnational parks to smaller corridors, they emphasized the importance of looking at connectivity for multiple species at multiple scales. They urged researchers and policy makers to take a more holistic multi-species approach to connectivity conservation.

Salvaging bycatch data for conservation: unexpected benefits of restored grasslands to amphibians in wetland buffer zones and ecological corridors, Mester et al. 2020

This study considers the effect of grassland restoration on amphibian populations in a 760-acre nature reserve – the Egyek-Pusztakócs Marsh System (EPMS) – established on former farmland in Hungary. The study shows that grassland restoration increased habitat range and quality for amphibians, extended hydrological supply, and limited genetic erosion among previously isolated populations. It also illustrates the role of smaller-scale ecological corridors.

Grassland restoration … creates corridors that maintain connectivity among the amphibian (sub)populations in the EPMS but it may also increase the permeability of the landscape to establish and maintain connections to other nearby metapopulations. Grassland restoration can thus also have an effect of minimizing genetic erosion of populations induced by isolation, which is one of the major causes of global amphibian decline [Mester 2020: 7].

Restoration can benefit amphibians by increasing the area of grasslands available for a variety of life activities such as foraging, burrowing, dispersal/ migration, or hiding from predators, aestivation and hibernation in the non-breeding period and by ensuring functional connectivity between wetlands both in the breeding and non-breeding periods [Mester 2020: 9].

This meta-analysis estimates total carbon sequestration potential in the US from various agroforestry practices to be 530 TgC/year (530 million metric tons), equivalent to about 1/3 of annual US carbon emissions from fossil fuel combustion. Based on their literature review, the authors estimate per-hectare sequestration rates (based on aboveground and belowground carbon accumulation) for each practice as follows: 6.1t C/ha/yr (silvopastoral), 3.4t C/ha/yr (alleycropping), 6.4t C/ha/yr (windbreaks), 2.6t C/ha/yr (riparian buffer).

Although often overlooked, ignored or taken for granted, earthworms are nevertheless keystone soil species, mediators and moderators for rebuilding healthy, biodiverse, high carbon and moisture rich topsoil [Darwin 1881; Blakemore 2016c]. We depend on soils for more than 99% of our food and 100% of our timber and natural fibres [Blakemore 2012, Pimentel 2013].  As an integral part of organic production, earthworms are key to agricultural sustainability and global ecosystem stability.  Ancient in origin (probably pre-Cambrian but certainly more than 500 million years old), the 7,000 known species of earthworms are ubiquitous and invariably associated with topsoil humus.  Earthworms are a basis of terrestrial food webs and the ultimate detritivor [Blakemore 2016c], recently reinstated as key players in the International “4 per 1000 Initiative” [4p1000.org, n.d.] to increase soil organic matter to store carbon. In this section, we discuss the abundance and variety of earthworms and their role in soil health and functionality.

Overview

Extrapolating data from Darwin [1881], their population numbers around 1.3 x 1015 or 1.3 quadrillion globally with biomass of 0.4 t/ha x 9.5 Gha of productive land = 3.8 Gt.  This is about ten times the biomass of all humanity, and twice that of both all domesticated stock and total global fish [Blakemore, 2017]. Forming possibly the largest beneficial animal resource on the planet, earthworms are yet apparently severely depleted by cultivation and agrichemical excesses of industrial farming, often being absent from such soils [e.g. Lee 1985] with both their populations and biodiversity in decline [Blakemore 2016a, b, c].

In comparison to intensive agrichemical farming, studies by Blakemore, [2000, 2016a, b] show a diverse array of up to 23 earthworm species per organic farm site (mean 13 spp), implicated in 16-80% increased crop or pasture yield (mean +39%) plus an average of 12% extra soil moisture storage (range 7-91%) compared to conventional neighbour farms. Carbon sequestration is restored at rates two to three times higher in pasture.  Such findings are highly relevant due to looming species extinction and climate change with requirement to meet the needs of a growing population.  Organic farming can thus produce higher yields and sequester more carbon.

Earthworms may number up to 1,000~2,000/m2 (10-20 million/ha, or 4-8 million/ac) in fertile soils with biomass as high as 3-5 t/ha, (1.2-2 t/ac ) so earthworm stocks may outweigh the above ground stock [Lee 1985; Blakemore 2016c, 2017].  Earthworm abundance and diversity increase in a truly sustainable system as they convert all organic ‘wastes’ into humus-rich compost while processing all atmospheric CO2 in 12 yr cycles [Blakemore 2016a].  Their burrows, as long as 9,000 km/ha (2250 mi/ac) [Kretzchmar 1982] and up to 15 m in depth (49.2 ft) [Sims & Gerard, 1999: 27, as cited in Blakemore 2016c] aerate, improve water infiltration and, importantly, provide habitats for many other beneficial organisms and microbes that they help distribute throughout the entire soil profile.  All rainfall is filtered through their burrows and water is stored in worm-worked humus.  Blakemore [2000] found up to 90% extra water in pasture compared to adjacent arable fields, and organic arable soil stored 40% more water than chemically farmed arable soil [Blakemore, Hochkirch 2017].

Wormless soils are economically and ecologically expensive: they need to be plowed regularly, and require extra irrigation plus subsidized artificial chemical nitrogen fertilizers and biocide sprays to fight off plant infections and infestations [Howard 1945; Balfour 1975].  This toxic burden has severe impact upon non-target organisms and any organism fed the crops – including humans – as well as poisoning the soil, air, waterways and oceans.  Such findings are summarized in Lady Eve Balfour’s IFOAM presentation in 1977 [Balfour 1977]. Another compelling reason for earthworm conservation is that it is impossible to “geoengineer” by addressing isolated variables the many benefits and essential irreducible systems services that earthworms freely and relentlessly provide.  In other words, we have no viable alternative to earthworms.

Soil and Earthworm Relationships

We face a complexity of inter-relating ecological problems.  Intensive chemical agriculture is a major GHG contributor (28-50%) and a major source of extraneous CO2 (currently 10-25% and in total historically up to 40%) [Houghton 2010: 338, 348]:

Globally, the conversion of lands to croplands has been responsible for the largest emissions of carbon from land-use change. . . From 1850 to 2000, land use and land-use change released an estimated 108–188 Gt (billion tons) of carbon to the atmosphere, or about 28–40% of total anthropogenic emissions of carbon (274 Gt C from fossil fuels) [Strassmann 2008].

The FAO [Gerber 2013] found that intensive industrial livestock farming (rather than organic husbandry) contributed 14.5% of human-induced GHG emissions.  A newspaper report [Bryce 2013]  comments:

The FAO’s last livestock report, a 2006 assessment titled Livestock’s Long Shadow, found that farms breeding chickens, pigs, and cows for meat and dairy products, produced a disconcerting 18% of global greenhouse gas emissions . . . Around 30% of global biodiversity loss can be attributed to livestock production, such as the spread of pasture land or turning over forests and savannahs.

Although these figures vary due to different formulas for budgeting, it’s clear that agriculture in all its forms, including the practice of forest clearance, is a major contributor to GHG emissions.

The traditional, innovative & scientific methods of non-chemical, organic farming and Permaculture appreciate the importance of earthworm conservation [Howard 1945; Balfour 1975; Mollison 1988].  As a key player in natural processes and crucial issues, Darwin’s “lowly earthworm”, although neglected, warrants re-ascendency to its former position as premier farm livestock [Howard 1945]. For our own health and for that of our planet, we urgently need wholly natural vermi-composting at all scales (from kitchen to continent) in order to replace synthetic fertilizers and to facilitate rapid transition to broad-acre organics that also has earthworm livestock at its core.   Enabling earthworms to restore healthy soils is vital to stabilizing climate. All organic ‘wastes’ and manures should be recycled via vermi-composting and appropriate management employed to enhance populations of field-working worms.

Earthworm Article Summaries

van Groenigen et al. 2014. In a recent meta-analysis, while not considering organic farming or carbon per se, this study confirmed earthworm presence corresponding to crop yield increases of 25%, which is comparable to average ~39% extra organic yield in soils with earthworm proliferations determined by Blakemore [2000, 2016b]. This supports earlier studies by Wollny [1890: Forschungen auf der Gebiet der Agrikultur-Physik, 13, s. 381] that found addition of earthworms to soil led to a marked increase of cereal grain by 35-50% and of straw by 40%.

Solomon 2013. 

Although earthworms are beneficial in gardens and agricultural fields, they are harmful to Michigan’s forests where they are an invasive species. . . . Earthworms are not native to Michigan and the Great Lakes region, at least not since before glaciers covered the region; they were brought here during European settlement in the 1800s or possibly earlier. Plants, wildlife and forests evolved without any of these creatures around. They are now an invasive species that harms forests.

Hardwood forests without earthworms have a thick layer of slowly decomposing leaves, or “duff” that promotes a rich community of wildflowers, tree seedlings and small animals. Earthworms change that environment dramatically by essentially consuming the duff, thereby destroying habitat and reducing fertility. In contrast to their effect in gardens, earthworms cause forest soils to become more compacted. As a result of habitat loss, fertility declines and soil compaction, these forests may be less productive and have poorer new tree regeneration in the long run.

Another view, from oligochaetologist (worm scientist) Rob Blakemore, is as follows:

Regarding popularized concerns about alien Asian invasive worms threatening to destroy American native forests, this may reasonably be regarded as part of a process that is commonly known as Ecological Succession [Odum 2005].

Ironically, the ecological concept of succession started with Thoreau and Cowles on studies of forest succession and on the Lake Michigan dunes.  Large parts of the northeastern North America were glaciated up to about 10,000 years ago completely destroying all land surfaces and forming the Great Lakes. When the ice retreated Nature returned in successive waves and, gradually, the soil, vegetation, and animals communities re-established and species continue to evolve.  

According to Darwin [1881] earthworms are supremely important for natural productivity and for the recent progress of human civilizations. In this context the woodlands of Michigan seem a relatively minor issue compared to species extinction and climate change.  Healthy soils generally harbour earthworms and it appears there had been insufficient time for these slow-moving and flightless organisms to colonize without fast-track via incidental intervention of most-recent human settlers, often as anglers on the Great Lakes.  

When exotic crops and stock were introduced around the world 10,000 years ago, so too were attendent earthworms and these have now spread to “pristine,” albeit transitional, woodlands.  The many benefits earthworms have for agricultural and unmanaged soils may cause some changes in more natural habitats but this is a virtually unavoidable and irreversible force majeure and fact-of-life.

Certainly there will be a new ecological balance in time, possibly at a different level of productivity and biodiversity.  That is the definition of succession.

Héry et al. 2008. Earthworms have been observed to increase methanotrophy (methane metabolic breakdown) in soil covering a landfill; this is most likely “due to the stimulation of bacterial growth or activity than to substantial shifts in the methanotroph community structure” [Hery 2008: 92].  

Earthworm-mediated bioturbation has been linked to an increase in methanotrophy in a landfill biocover soil (AC Singer et al., unpublished), but the mechanism of this trophic interaction remains unclear. The aims of this study were to determine the composition of the active methanotroph community and to investigate the interactions between earthworms and bacteria in this landfill biocover soil where the methane oxidation activity was significantly increased by the earthworms [Hery 2008: 92].

And

We proposed the hypothesis that earthworms could stimulate the growth or the activity of methanotrophs. We showed that the earthworm-mediated increase of methane oxidation in the landfill biocover soil only weakly correlated with a shift in the structure of the active methanotroph population. Future work needs to focus on the relationship between this earthworm effect on enhanced methane oxidation in landfill cover soil and this effect on bacterial activity and growth. The possible contribution of an enriched population of nitrifying bacteria to methane oxidation also requires further investigation [Hery 2008: 101].

Grasslands have been estimated to cover approximately 40% of global land surface area, approximately 5.25 bn ha (~13 bn ac ), except for Greenland and Antarctica [Suttie 2005; White 2000:12].  Their deep soils are rich repositories of nutrients, especially carbon and water.  Many grasslands are anthropogenic, i.e., resulting from various land management techniques to maintain land for grazing and crop production by humans.  Virgin grasslands are increasingly rare, possibly leading to significant underestimations of their potential positive contribution to productivity, and to carbon and water storage.  Grasslands are important repositories of biodiversity, and have significant impacts on weather and climate.  Here we review research and articles that indicate soil carbon storage potentials of roughly 13 gigatons per year (the equivalent of 6.5 ppm) if global grasslands were managed regeneratively.

Overview

While we have separate sections for Grasslands, Croplands and Soils, there is inevitable overlap. Many croplands are modified grasslands, and both are, of course, based in soils.  Yet there are enough differences in each area of study to merit separate sections, keeping in mind that systemic behaviors and interactions are broadly applicable.

Since the onset of agriculture over 10,000 years ago with land management techniques that expose soil to air, estimates of up to 537 gigatons of soil carbon have been oxidized to carbon dioxide and other greenhouse gases [Buringh 1984: 91].   Even so, soils (>2,300 Gt) currently hold almost as much carbon as plants (550 Gt), atmosphere (800 Gt) and ocean surface waters (1,000 Gt) combined [NASA 2011], and almost surely retain the potential to store enough atmospheric carbon to return to pre-industrial levels.  

Typical soil studies examine the first 30 cm (1 ft) of soil depth, but more recent investigations indicate that major soil carbon storage takes place deeper than that, often in a more stable form [Liebig 2008, Follett 2012, Harper 2013].  A USDA paper found unexpectedly high quantities of soil organic carbon (SOC) between 30-150 cm (1-5 ft) below the surface, exceeding 2.25 tC/ha/yr (0.9 tC/ac/yr) [Follett 2012]. A study of switchgrass for bioenergy found rates of SOC increase of up to 2.75 tC/ha/yr (1.1 tC/ac/yr) when measured to depths of up to 120 cm (4 ft) [Liebig 2008].  On an intensively grazed former row-crop agriculture land converted to dairy farms in the Southeastern U.S., Machmuller et al. found many improvements in the sandy soil, including ~1.25 tC/ha/yr (~0.5 tC/ac/yr) sequestration after accounting for ruminant methane emissions [Machmuller 2015].  In addition, the ultimate methane emissions may have been markedly less than measured since the report did not consider methane breakdown into CO2 from methanotrophic bacteria and atmospheric hydroxyl radical oxidation, with a significant reduction of methane’s ultimate greenhouse gas impacts.  

These reports demonstrate the potential for massive amounts of soil carbon storage, significant cooling of the biosphere, and dramatic improvements in ecosystem health using regenerative approaches to grassland management.

Grassland Evolution

Grasslands have long been a rich repository of carbon, both stable and labile. The co-evolution of grasslands with grazing ruminants has contributed to dramatic global cooling over the past 50 million years as a result of significant photosynthetic carbon drawdown into grassland soils [Retallack 2013].  Thus, grasslands are more than a consequence of geophysical changes, they are

. . . a biological force in their own right (Retallack 1998), in some ways comparable to the human rise to dominance of planetary resources (Vitousek et al. 1997). Grasslands have long been considered products of the coevolution of grasses and grazers (Kovalevsky 1873). Few plants other than grasses can withstand the high-crowned, enamel-edged teeth and hard hooves of antelope and horses. Yet these same animals are best suited to the abrasive gritty opal phytoliths and dust of flat, open grasslands. Grasses recover readily from fire and nurture large herbivores such as elephants: both fire and elephants promote grassland at the expense of wood land (Retallack 1997b; Jacobs et al. 1999). Grasses suppress insect and fungal attack with secondary metabolites such as cyclic hydroxamic acids (Frey et al. 1997). Grasses create Mollisols, unique soils with fine crumb clods rich in organic matter (Retallack 1997b).  (Retallack 2001:407) [Emphasis added.]

Occupying such vast areas of planetary land surface, grasslands have a major influence on the global climate:

CO2 and CH4 (which rapidly oxidizes to CO2 ) are important greenhouse gases, and mechanisms for burial of their C may result in climatic cooling (Berner 1999). The most important long-term C sink from grasslands is their supply by erosion to sedimentary basins of crumb peds, which are unusually rich in organic matter intimately admixed with clay (Pawluk and Bal 1985). Tropical forests, in contrast, yield highly oxidized spherical micropeds with virtually no organic content (Retallack 1991a).” (Retallack 2001:414)

While there are other soils that are greater carbon sinks, such as peat bogs, wetlands and coastal habitats (e.g., mangroves, seagrasses), for volume and depth of carbon storage on vast areas of land, grasslands have enormous potential:

Grasses themselves are C sinks, especially considering their mass of roots and rhizomes underground. . . . [Numerous soil investigations] indicate that past estimates of organic C in tropical grassland soils have been low, in part because soils were not analyzed to sufficient depths. . . . Grassland and woodland soils may have comparable amounts of organic C in the surface 15cm. Beyond that depth, organic C values drop off dramatically in woodland soils but remain high in grassland soils to a meter or more. The fine structure and fertility of grassland soils is in large part due to this large C reservoir. (Retallack 2001:415)

Conventionally, it is estimated that approximately 40% of global land surface area is grasslands (52.5 million square kilometers, or ~5.25 billion hectares, or ~13 billion acres [Suttie 2005]), except for Greenland and Antarctica [White 2000:12; see Figure 1, below].  This is likely a significant underestimation of soil surface area and volume, since grasslands are not uniformly flat, with topographical variations adding carbon, water, etc. to areas that are typically calculated on the basis of a two-dimensional map projection [Blakemore 2016: fig. 5]. The implications are that there may be considerably greater volumes of soil amenable to regeneration, carbon capture and water storage than is conventionally assumed (see the section, “Do We Have More Soil for Carbon Storage than We Thought?”).

Natural grasslands are typically areas of low and seasonal rainfall. Unlike temperate environments with year-round rainfall, semi-arid and arid grasslands are dependent on grazing animals as a keystone species. The habits of grassland plants are as dependent on grazing animals as the animals are dependent on the plants that grasslands provide as food.

                          Figure 1: Global Extent of Grassland (White 2000:12).

Grazing animals are ruminants and as long as they are grazed in herds that move frequently, as they do in natural habitats in the company of predators, ruminant species are often interchangeable insofar as grassland health is concerned. Bison and antelope, for example, may be the wild ruminants that roamed the prairies and savannahs, but domestic cattle, when properly managed (human herders are the equivalent of wild predators), serve the same ecosystem functions.

Pioneering Work Of Allan Savory

Allan Savory, a wildlife biologist from Zimbabwe, began studying desertification in the 1950s.  He pioneered an approach that he has termed Holistic Planned Grazing (HPG) for regenerative management of grasslands. He noted that there are essential differences between temperate grasslands, which he termed “non-brittle” environments, and arid and semi-arid grasslands, which he referred to as “brittle.”  These distinctions are critical in understanding how different habitats require different management approaches.

Non-brittle environments, because of year-round rainfall, are relatively forgiving of mismanagement that destroys soil biota and exposes soil to sunlight, air and the elements. Recovery from soil degradation can be be rapid.  Brittle environments, to the contrary, are fragile and easily desiccated, and when poorly managed, either from overgrazing or undergrazing, may take decades or centuries to recover or even ultimately turn to desert.

Brittle environments particularly need ruminant evolutionary partners. Their hooves are designed to open the soils to air and water, and their digestive systems deposit a feast for soil organisms.  The ruminant gut is a moist refuge during the dry season for soil microbes which are essential to the health of the land. Grasses need to be bitten lest they shade out their own new growth.

What Savory discovered is that the same land may either flourish or die depending on how it is grazed.  When ruminants are kept in check by predators they graze an area in tight herds for protection and then move to the next patch of fresh grasses and other plants, providing up to two years of recovery and regrowth time for the recently grazed pasture. On the other hand, when they are provided the safety of fencing and left to graze large areas at will, they return to their favorite plants and overgraze those areas, eventually compacting the soil, preventing water infiltration and proper aeration, killing the plants, and leading to desertification.

The difference in land health is dramatic.^[19]  The pictures below illustrate:

Mexico

Arizona

Zimbabwe

Fig. 1. These pictures are of neighboring properties in Mexico, Arizona and Zimbabwe.  In each area they were taken on the same day, have similar soils, and the same precipitation.   The pictures on the right are examples of properly managed livestock through Holistic Planned Grazing to restore grasslands.  On the left we see examples of improperly managed livestock as well as exclusion from grazing (“resting the land”) [Savory Institute 2015:12]

Savory’s work, after decades of successful application on ranches in Africa, Asia, Australia and North and South America, garnered global attention (and controversy) after his TED Talk in 2013 [Savory 2013].  

The primary point is that If well-managed, grassland soils can not only sequester annual greenhouse gas emissions but can also begin to draw down legacy atmospheric carbon as well.  In addition, they provide human and other predator food, converting grasses inedible to non-ruminant mammals to meat. 

Grasslands As Ecosystems

It is useful to understand how grasslands work as intact ecosystems, thereby providing a solid theoretical basis for observations of grasslands as vast carbon and water sinks.  What appears above-ground is only a hint of grassland ecosystem dynamics.  Soils are the planet’s most complex and least understood terrestrial ecosystem, yet soils are where most of the action takes place on grasslands as well. While all soils are built on the foundation of minerals provided by weathering of rock, the soils on grasslands are primarily biological soils.  The kingdoms of life are the active agents in soil creation and it is the interactions among life forms that create the rich and productive grassland soils (see Soils section).

The basis for all life is the microbial kingdom.  These smallest of cells, with their complex biochemistry, morphology and behavior, are active players in creating stable soil molecules, storing abundant carbon and water. In addition to microbes, soil ecosystems are built from exchanges among fungi, insects, worms, green plants, birds and small and large mammals. It is this set of rich interactions that creates the biodiverse, abundant and resilient environment of global grasslands.

For the sake of illustration, let us start the discussion with the ruminant gut during a dry season.  Microbes survive in a warm, moist environment while constantly being cycled into the soil through digestion and elimination.  The short-term hoof disturbance with minimal compaction while animals are constantly moving opens the soils to available moisture from precipitation, urination and condensation.  During the rainy season the water is more effectively absorbed into opened soils, nourishing plants, raising the water table and eventually even leading to perennial streams and ponds. Even limited rainfall goes a long way in spongy soils that are covered with grasses and other plants to keep the ground cool and moist [Byck 2014: 8’38”]

Methane

Methane, a relatively short-lived but powerful greenhouse gas, is often raised as a serious concern with beef production.  This is surely true when animals are left to roam freely and overgraze, and then moved to concentrated animal feeding operations with large manure lagoons.  However, it is important to consider the whole of ecosystem functions in assessing methane emissions {Savory Institute 2015].  This includes the conversion of the methane molecule into carbon dioxide by bacteria (metanotrophs) that live in healthy soils and literally eat energy-rich methane, and the oxidation of methane by hydroxyl radicals present in the lower atmosphere. It may include other ecosystem processes, such as the effects of earthworms increasing methanotrophic bacterial activity in landfills and pastures [Héry 2008; Kernecker 2014]. The result is a virtuous cycle where plants can then take up that carbon dioxide through photosynthesis and send some of the carbon back underground through their root systems.  Throughout their life cycle under conventional industrial management, cattle are rarely if ever exposed to such healthy, biodiverse soils.  

Historical methane data indicates that in the United States, for example, pre-settlement wild ruminants generated approximately 86% of the methane of current farmed ruminants (Hristov 2012:1371).  Yet even with vast numbers of ruminants on grasslands across the planet, atmospheric methane remained constant until the global dependence on widespread use of fossil fuels and its effects on agriculture and animal husbandry began to grow rapidly in the 18th century (Fig. 2).

Thus, results are very different with animals grazed in a manner that mimics nature.  Rowntree et al. describe the importance of accounting for the beneficial ecosystem services that well-managed grazing systems can provide.

. . . LCA’s [Life Cycle Assessments] often consider soil C to be in dynamic equilibrium. However, empirical data suggest otherwise (e.g. Machmuller et al., 2015; Teague et al., 2011). Recent studies such as Ripple et al. (2014) and Eshel et al. (2014) have reported the emissions from ruminants in food production without accounting for the beneficial ecosystem services that well-managed grazing systems can provide. In our study, we used 3 tC/ha/yr (1.2 tC/ac/yr) as a potential C sequestration figure, which is relatively high (Conant et al., 2001) but viable based on existing studies (Teague et al., 2011; Delgado et al., 2011; Machmuller et al., 2015; Teague et al., 2016). Importantly, the results presented here suggest that with appropriately managed grazing, a grass-finished beef model can not only contribute to food provisioning but also be ecologically regenerative as well. [Rowntree 2016:36]

This excerpt illustrates a paradigm shift in action, wherein investigators within the dominant paradigm are constrained from evaluating the possibilities offered by “beneficial ecosystem services” because it doesn’t occur to them to consider them.  Such biological processes are invisible due to limiting assumptions of the paradigm.

Research into systems implications of holistically managed grass-finished beef is growing.  It is only recently that mainstream researchers are beginning to understand that the biological function of animals in an ecosystem is as dependent on the ecosystem as it is on the biology of the animal.^[20]  Studying animals in isolation or as part of a synthetic system such as industrial agriculture often leads to incorrect conclusions.

Fig. 2.  Historical Concentrations of Greenhouse Gases [IPPC 2007]

Review of some studies of grasslands

Soil creation (pedogenesis) is conventionally defined as the weathering of rock; it may take 3,000 years or longer to create a foot of soil through geological processes.  Soil created through biological activity, on the other hand, happens orders of magnitude faster, up to several inches per year.  

Australian soil scientist Christine Jones notes that

The rates of soil formation provided in the scientific literature usually refer to the weathering of parent material and the differentiation of soil profiles. These are extremely slow processes, sometimes taking thousands of years. Topsoil formation is different and can occur rapidly under appropriate conditions. . . .

The late P.A. Yeomans, developer of the Keyline system of land management, recognised that the sustainability of the whole farm was dependent on living, vibrant topsoil. The formation of new topsoil using Keyline principles, at rates not previously considered possible, was due to the use of a tillage implement designed to increase soil oxygen and moisture levels, combined with a rest/recovery form of grazing and pasture slashing, to prune grass roots and feed soil biota, especially endemic earthworms. Yeomans was able to produce 10 cm of friable black soil within three years, on what was previously bare weathered red shale on his North Richmond farm (Hill 2002).

Bennett (1939) calculated a rate of topsoil formation of just over 11 t/ha/yr (4.4 t/ac/yr) for soils in which organic material was intermixed into surface layers. In situations where plant root mass is high, rates of topsoil formation of 15-20 t/ha/yr (6-8 t/ha/yr) have been indicated (Brady 1984). Healthy groundcover, high root biomass and high levels of associated microbial activity, are fundamental to the success of any technique for building new topsoil.

If the land management is appropriate, evidence of new topsoil formation can be seen within 12 months, with quite dramatic effects often observed within three years. Many people have built new topsoil in their vegetable or flower gardens. Some have started to build new topsoil on their farms. If you have not seen new soil being formed, make a point of doing so. (Jones 2003:19-20)

Healthy biodiverse grasslands with abundant animal populations provide favorable circumstances for biological soil accumulation and carbon sequestration, including opening soils to air and water, fertilizing soil life and stimulating growth of grasses.  

Fig. 2. “Root Systems of Prairie Plants,” Heidi Natura, Conservation Research Institute, n.d., http://kmlandtrust.org/pdf/NPGpp5-6-11×17.pdf

Using current best land management practices, recent research has begun to confirm the importance of studying soil organic carbon accumulation on grasslands.  In a 9-year study of bioenergy crops, investigators found that switchgrass and maize stored 50% of their soil organic carbon (SOC) below 30 cm (1 foot), up to 4 times more than used in models in use at that time (Follett 2012:866):

Most of the research on SOC in agricultural production systems focused on C in the 0 to 30 cm depth [22–27]. A few studies in which soil sampling has been conducted at greater depths indicate that production agriculture affects soil C deeper in the soil profile [28,29]. (Follett 2012:867)

Concerns about soil depth measurement are not new [Liebig 2008]; depth of soil measurement in estimating soil carbon storage potential is a significant issue.  Conventional soil science, which largely addresses agricultural soils managed in industrial agricultural contexts, typically measures soil carbon down to around 30 – 40 cm (12 – 16 inches).  Yet roots of native prairie plants may reach 5 times that depth (see Fig. 1), storing carbon in stable molecules for centuries and millennia as long as the soils are undisturbed and not exposed to light or desiccation.  [See Soils section.]

Methodological issues for assessment of SOC have thus been problematic, and have likely led to serious overall underestimation of soil sequestration capacity. This is particularly relevant because these soils will not be able play their appropriate critical role in addressing climate until mainstream science and policy recognize and promote the potential of best practices in land management in all ecosystems.  Harper & Tibbett found up to five times more soil carbon in Australian soils at depths greater than 1 meter (~3 feet)  than is conventionally estimated:

When the SOC storage within the deep profiles was compared with what would have been reported from conventional sampling depths (Table 1), it is clear that considerably more SOC was stored in the soils than is normally reported. Across all samples, the surface 0.5 m, which is deeper than the standard IPCC sampling depth of 0.3 m (Aalde et al. 2006), contained 5.8± 0.57 kgCm −2 or 21 % of the total store to bedrock. [Harper 2013: 645]

We discuss the dynamics of water cycling and forests elsewhere, but it is worth noting a recent paper suggesting that a more holistic view of ecosystem dynamics is in order. A paradigm shift prioritizing water over carbon as the driving climate force more accurately and effectively guides climate recovery strategies and offers more tactical and regenerative options [Ellison 2017; s.a. Schmidt 2017].  Water is more tangible to most people than carbon, and water recovery is more visible and rapid, offering hope and encouragement in a generally grim scenario.  In addition there are numerous other benefits to people and landscapes with improved water management.

THE IMPORTANCE OF ANECDOTAL EVIDENCE

While anecdotal evidence is often disdained in academic science, in many scientific pursuits it forms the foundation of future study. This is especially true in such predominantly observational pursuits such as naturalist biology, ecology, rangeland science and agronomy, where study of the visible vagaries of the natural world inspires questions that may reach far beyond what a happenstance occurrence would imply.  It is the weight of such observations that leads to formal protocols, hypotheses and theories to explore the details of a field.  Isolating variables can be a very helpful tool; unfortunately, modern science practice has generally lost sight of systems contexts, and that system behavior can be very different from the behavior of any of its isolated parts.

Therefore, it is essential to embrace both analytical and holistic evidence in order to build a full understanding of how environmental systems work as wholes. Together these complementary approaches provide a more comprehensive picture of the systems in question, as well as much clearer guidance for how to proceed in current global ecosystem urgencies.  

A small selection of a growing literature of informative anecdotal reports of grassland eco-restoration is included among formal studies below.  They reflect the extensive experience of farmers, ranchers and other land managers, and demonstrate the potential positive effects of regenerative land management and eco-restoration on climate and the biosphere as a whole. See for example Stigge 2016, Oppenheimer 2015, Byck 2014, Brown n.d., Brown 2016.

Grassland Article Summaries

Byck 2014. This 12-minute video relates the experiences of three ranchers who manage cattle and land according to regenerative land management principles. They discuss their transition to Holistic Planned Grazing, where for two of them, in areas of ~15 inches of rainfall, their alternative had been bankruptcy. The video vividly illustrates the benefits of cover-cropping and organics, the improvement in lifestyle and economics, and the dramatic improvement in biodiversity and water management.  [Byck 2014]

Follett 2012.  A USDA study found unexpectedly high quantities of soil organic carbon (SOC) between 30-150 cm (1-5 ft) below the surface, exceeding 2.25 tC/ha/yr (0.9 tC/ac/yr). Ausmus reports that

. . . a 9-year project that evaluated the effects of nitrogen fertilizer and harvest treatments on soil organic carbon sequestration in switchgrass and no-till maize crops managed for biofeedstock production [found that] more than 50 percent of the soil carbon was found between 1 and 5 feet below the soil surface. The average annual increase of soil organic carbon throughout the first 5 feet of subsoil also exceeded 0.9 tons per acre per year [Ausmus 2014: 4-5].

Of interest were the difficulties the authors faced in getting the study through peer review and published since their results were so unexpected, as reported in Ausmus 2014.^[21]  It originally appeared in Bioenergy Research in 2012. [Follett 2012].

Note too that the Follett study was performed on already degraded soils, “Perennial grasses could be used as bioenergy crops on about 20 million ha (ha = 10,000 m2 or 2.5 acres) of marginal or idle cropland in the USA alone [18]” (Follett 2012:867). [Emphasis added.]  A well-supported inference is that healthy, biodiverse soils will yield even better results.

Degraded soils may be less effective carbon sinks than virgin soils even though they have lost most of their carbon because the soil life that creates long-lasting stable carbon molecules is damaged or destroyed by synthetic inputs, tilling and other forms of mismanagement. Nonetheless, results were dramatic:

In the first 9 years of a long-term C sequestration study in eastern Nebraska, USA, switchgrass and maize with best management practices had average annual increases in SOC per hectare that exceed 2 tC/yr for the 0 to 150 cm soil depth. For both switchgrass and maize, over 50 % of the increase in SOC was below the 30 cm depth. SOC sequestration by switchgrass was twofold to fourfold greater than that used in models to date which also assumed no SOC sequestration by maize. (Follett 2012:866) . . . .

Our results clearly show that significant amounts of C were sequestered deep in the soil profile by switchgrass grown and managed as a biomass energy crop and maize grown continuously in a no-tillage production system for the cultivar Trailblazer array of N fertility and harvest treatments for a 9-year period. For almost all other C sequestration reports used in bioenergy models, studies designed for other purposes were adapted to obtain soil C sequestration estimates, initial soil samples were not available, and control samples were from adjacent fields or non-treatment areas. Our results are supported by similar results reported by Liebig et al. [28] for four switchgrass fields managed with uniform N rates and harvest treatments for 5 years in the USA western Corn Belt. They are supported by the recent work of Varvel and Wilhelm [29] for maize in which significant increases in soil C occurred in soil layers up to 150 cm in depth in maize no-till plots as compared to tilled plots. The soil C that is sequestered deeper than 30 cm is expected to be more stable over time since it is below the tillage zone. Even in the top 30 cm of soil, sequestered C may be stable for extended periods in no-till production systems as reported previously by Follett et al. [33].  (Follett 2012:873)

Harper 2013. This study suggests that the standard sampling depth of 30cm vastly underestimates the global store of soil organic carbon, and thus, presumably, the potential of future soil carbon storage that could result from eco-restoration efforts. “Hypothesizing that SOC retained in soils below the top half metre (in highly weathered deep profiles) would account for the major proportion of SOC in the landscape” [Harper 2013: 642], this southwestern Australia study took deep soil samples in 38 spots across 5 locations. They found that 79% of total carbon store to bedrock occurred below a half meter depth, and 41% occurred below 5 meters depth. “There are two possible sources for the deep carbon; that produced in situ by roots or dissolved carbon that has moved downward from nearer the surface” [Harper 2013: 645]. The sample sites were recently reforested or were under agriculture, and previously had been “covered in a range of xerophytic plants, with root systems that extended to depths of 40 m, such as reported for a Eucalyptus marginata forest” [Harper 2013: 642]. The authors note that more research is needed to understand how deep SOC is affected by land-used changes and climate change.

When the SOC storage within the deep profiles was compared with what would have been reported from conventional sampling depths, it is clear that considerably more SOC was stored in the soils than is normally reported. Across all samples, the surface 0.5m, which is deeper than the standard IPCC sampling depth of 0.3 m (Aalde et al. 2006), contained 5.8± 0.57 kgCm−2 or 21 % of the total store to bedrock. If this is adjusted to 0.3 m depth, using an exponential function based on the samples in the surface metre, the value decreases to 5.6 kgCm−2. For the individual sites this ranged from 3.6 to 8.0 kgCm−2, or 14 –37 % of the total store. …the surface 5 m contained 16.3±1.38 kgCm−2 or 59 % of the total store to bedrock, with this proportion varying from 47 to 77% across the five sampling locations. The amount of carbon stored in the soils can also be contrasted with the biomass carbon storage of 11.0–16.0kg Cm−2 expected at equilibrium following reforestation for these sites (Harper et al. 2007) and likely previously removed from the sites by deforestation in advance of agriculture [p.645].

Liebig 2008.  A study of switchgrass for bioenergy found rates of SOC (Soil Organic Carbon) increase of up to 2.75 tC/ha/yr (1.1 tC/ac/yr) when measured to depths of up to 120 cm (4 ft).  “In this study, switchgrass significantly affected change in SOC. . . Across sites, SOC increased significantly at 0–30 cm (1 ft) and 0–120 cm (4 ft), with accrual rates of 1.1 and 2.9 tC/ha (0.44 and 1.16 tC/ac), respectively.” [Liebig 2008:215]  This indicates the chronic underestimation of soil carbon capacity in the many studies which by convention only measure SOC down to 30 cm (1 ft).

Machmuller 2015.  On intensively grazed former row-crop agricultural land converted to dairy farms in the Southeastern U.S., Machmuller et al. found many improvements in the sandy soil, including ~1.25 tC/ha/yr (~0.5 tC/ac/yr) sequestration after accounting for ruminant methane emissions.  The study “sought to determine how fast and how much soil C accumulates following conversion of row crop agriculture to management-intensive grazed pastures in the southeastern United States. . . .  The highest rates of belowground C accumulation occur when land is converted to grassland ecosystems” [Machmuller 2015: 2]. These intensively grazed managed systems led to an approximately 75% increase in soil carbon within six years,

[a] high C accumulation rate [that] stems from year round intensive forage/grazing management techniques on sandy soils with an initially low soil C content due to past conventional-till row crop agriculture. . . . These forage-management techniques are precisely those suggested to increase SOM in pasture systems and when they are applied to soils with degraded SOC content, such as soils in the southeastern United States, rapid C accumulation ensues. . . .

On the basis of a whole farm C-cycle analysis, C accumulation appears to offset methane emissions during the rapid soil C accumulation phase . . . As the C accumulation rate declines these farms will become net C-emitting—similar to all dairy production—because of ruminant methane emissions.  However, the substantial soil-quality benefits of higher organic matter remain and will likely increase the sustainability of dairy production using management-intensive grazing [Machmuller 2015:3].

The eventual methane emissions may be markedly less than suggested, however, since the report did not consider methane breakdown from methanotrophic bacteria and atmospheric hydroxyl radical oxidation.  

The authors conclude

that pasture-based intensively grazed dairy systems may provide a near-term solution for agricultural lands that have experienced soil-C loss from previous management practices. Emerging land uses, such as management-intensive grazing, offer profitable and sustainable solutions to our needs for pairing food production with soil restoration and C sequestration. [Machmuller 2015: 2-3]

McCosker 2000.  A discussion of the introduction of what the author calls “cell grazing,” framed as a paradigm shift in Thomas Kuhn’s terms [Kuhn 1962] over the years 1990-1999.  McCosker reviews the dichotomy between researchers and producers, and travels to see actual results in the U.S., Zimbabwe, Namibia, and South Africa. He states, “Only after seeing the outcomes time and again in all possible environments was I finally convinced that the principles could not be faulted.” This kind of paradigm shift has been experienced repeatedly as practitioners must develop the courage to challenge prevailing assumptions in order to take the requisite transitional steps.

Oppenheimer 2015.  Farmers Paul and Elizabeth Kaiser own eight acres in Sebastopol in Northern California and they farm three of them, developing a model that other farmers are beginning to use.  

[Kaiser] farms a mere eight acres, and harvests fewer than three of them. Nonetheless, his methods are at the forefront of a farming movement that is so new (at least in the U.S.), and so built for a climate-changed world of diminishing rains, that it opens up gargantuan possibilities. One might call this methodology sustainability on steroids, because it can generate substantial profits. Last year, Kaiser’s Sonoma County farm grossed more than $100,000 an acre, which is 10 times the average per-acre income of comparable California farms. This includes Sonoma’s legendary vineyards, which have been overtaking farmland for decades, largely because wine grapes have become much more lucrative these days than food, at least the way most farmers grow it.

Kaiser manages all of this without plowing an inch of his ground, without doing any weeding, and without using any sprays—either chemical or organic. And while most farmers, even on model organic farms, constantly tinker with various fertilizer cocktails, Kaiser concentrates on just one: a pile of rotten food and plants, commonly known as compost, and lots of it. Kaiser then adds this compost to a rare blend of farming practices, both old and new, all aimed at returning dirt to the richest, most fertile seedbed possible.

They use permaculture, agroforestry and other intensive techniques, have built deep, healthy soils by keeping the ground covered and spongy to capture water and carbon, all of which provide solid protection from droughts and floods.

Retallack 2001, 2013.  “Grassland expansion initiated increased organic C storage in soils, soil water retention, speed of nutrient exploitation, surface albedo, and C burial in sediments eroded from their soils. These changes had many consequences, including long-term global cooling.” [Retallack 2001:422] and “This climatic zone is not only the most widespread, but also the most fertile region of our planet.” [Retallack 2013:78]  The paleohistory of grasslands provides the basis for considering the potential of grasslands as huge biological carbon sinks that may be realized again with regenerative land management

Rodale Institute 2014. 

Rodale reports that regenerative grazing practiced on a global scale could sequester 71% of annual emissions of 14 Gt C/yr.  Combined results from regenerative grazing and agriculture techniques could, if practiced globally, lead to a net reduction of atmospheric carbon dioxide of 1.7 gigatons per year, or 0.85 ppm/year. This represents a potential sequestration of approx. 3t C/ha/yr (1.2t C/ac/yr) on grasslands and croplands. [Rodale Institute 2014:9]  This does not include the considerable contributions of non-agricultural lands, nor recent developments in intensive regenerative practices such as permaculture and biochar.

Moreover, Rodale’s side-by-side trial after 30+ years showed that, following the three-year transition period, organic yields match conventional yields, except in drought years, when organic yields surpass conventional yields. Furthermore, energy input and greenhouse gas emissions were lower in organic systems, and profits were higher.

Rowntree 2016.  Examining ruminant methane and net carbon sequestration for grassfed beef in a systems context, Rowntree found net sequestration rates of up to 2.11 tons/ha/yr (0.84 tC/ac/yr) for non-irrigated, lightly stocked grazing.

Therefore, including soil carbon sequestration (SCS) potential brings the differences in grazing environments into focus, and significantly changes the outcome.  When comparing two well-managed grazing strategies, grass-finished (MOB) and conventional (IRG), each strategy could be an overall carbon sink, but the MOB grazing would only need sequester half as much carbon (1 tC/ha/yr) as IRG grazing  (2 tC/ha/yr) for a net zero greenhouse gas footprint.  Methane emissions were similar in both grazing environments, but MOB grazing offered significant benefits in increased carbon sequestration. [Rowntree 2016:36]

It is unlikely that such SCS would take place in the absence of a healthy biodiverse ecosystem, one that is supported through MOB grazing. The result is higher net methane emissions under conventional grazing practices. The opposite occurs with properly managed grazing practices, where grasslands as a system actively build more soil carbon for years, leading to a net increase in soil carbon despite enteric methane production by ruminants.^[22] 

The recent call for improved management of grazing systems as part of an international climate change mitigation strategy is critical, particularly in light of many existing beef LCAs [Life Cycle Assessments] that have concluded that beef cattle produced in grazing systems are a particularly large sources of GHG emissions. To identify the best opportunities to reduce GHG emissions from beef production, a systems approach that considers the potential to increase soil C and reduce ecosystem-level GHG emissions is essential… [W]e generated an LCA that indicates highly-managed grass-finished beef systems in the Upper Midwestern United States can mitigate GHG emissions through SCS while contributing to food provisioning at stocking rates as high as 2.5 Animal Units (AU) per hectare. From this data, we conclude that well-managed grazing and grass-finishing systems in environmentally appropriate settings can positively contribute to reducing the carbon footprint of beef cattle, while lowering overall atmospheric CO2 concentrations. [Rowntree 2016:36]

Schwartz 2013.  Cows Save the Planet was a landmark book that helped launch the regenerative agriculture and eco-restoration climate movements. The author gathers stories from practitioners around the world and paints a picture of broad possibilities for addressing global warming, floods, droughts, desertification, malnutrition and many other seemingly unrelated problems that have a single common cause: human mismanagement of lands across the planet. An excellent introduction for a general readership as well as for scientists unfamiliar with the potentials that nature provides.

Shinn n.d. Ridge Shinn is a rancher in Central Massachusetts (Big Picture Beef) who introduced Allan Savory’s methods to the state.  For over ten years he has been investigating the holistic context for bringing healthy meat through the local supply chain, and its relationship to human health and global warming.  He states,

Since the 1990’s, science has discovered important connections between rotational grazing, soil health, and healthy food. Big Picture Beef’s methods for raising 100% grass-fed cattle offer huge benefits for the environment and for society.  The long term goal of the program is Northeast beef for Northeast markets, carbon sequestration, soil fertility and biodiversity, energy savings, and a revitalized rural economy.

Teague et al. 2016. In a review of the literature, the authors conclude that regenerative conservation cropping and adaptive multi-paddock grazing can turn agricultural soils from a carbon source in conventional agriculture into a carbon sink at rate of ~3 tC/ha/yr (~1.2 tC/ac/yr).  Key factors include the use of no-till, cover crops, managed grazing, organic soil amendments and biotic fertilizer formulations. These practices can result in elimination of soil erosion and loss, the greatest agricultural contribution to global warming (1 Gt C/yr).  Benefits may include “increased water infiltration, improved water catchment, greater biodiversity, increased ecosystem stability and resilience, and improved C sequestration.” [Teague 2016:158]

Conclusion

These reports demonstrate the promising potential for storage of massive amounts of soil carbon to address both the atmospheric and eco-destruction aspects of climate, along with dramatic improvements in ecosystem health using regenerative approaches to grassland management.

Note: As mentioned in the Release notes, we have a small staff, and therefore have had to postpone some important material for the next release, scheduled for January 2018.  This is particularly true of forests and we will include a more thorough examination of their importance in addressing climate moving forward.  Nonetheless, we felt that the investigations here were innovative and interesting, and we wanted to make them available to our readers sooner rather than later.

Forests cover nearly 31% of Earth’s total land area [FAO 2016], and remain one of the major terrestrial ecosystems on the planet. Forests play a significant role in the global ecosystem through cooling, evapotranspiration, covering/shading/sheltering, providing fuel and fiber, aiding cloud formation, and creating wind. Because global forests and wooded lands store an estimated 485 Gt of carbon, forest conservation and afforestation are recognized in the United Nations Framework Convention on Climate Change (UNFCCC) as key strategies for climate change mitigation [UNFCCC 2017].  

Despite this acknowledgment, “for the world as a whole, carbon stocks in forest biomass decreased by an estimated 0.22 Gt annually during the period 2011–2015. This was mainly because of a reduction in the global forest area”  [UNFCCC 2017]. Indeed, humanity has been in the business of clearing forests for thousands of years, and this continues today. However, rapid reductions in deforestation could abate further carbon emissions and thus extreme results of climate change. Moreover, reductions in deforestation and implementation of agroforestry practices together could restore biodiversity in damaged ecosystems, repair local and global water cycles, and, ultimately, help restore carbon levels to pre-industrial levels. Here we present several articles illustrating the impact of forests on global climate, as well as the potential for restorative afforestation and agroforestry practices to sequester large amounts of carbon.

Forest Article Summaries

Ellison 2017.  This paper takes the innovative and paradigm-shifting position that carbon is not the primary consideration in climate; rather, it is water that should be a central focus in assessing climate processes and effects. It considers forests from a systems perspective.

Forest-driven water and energy cycles are poorly integrated into regional, national, continental and global decision-making on climate change adaptation, mitigation, land use and water management. This constrains humanity’s ability to protect our planet’s climate and life-sustaining functions. The substantial body of research we review reveals that forest, water and energy interactions provide the foundations for carbon storage, for cooling terrestrial surfaces and for distributing water resources. Forests and trees must be recognized as prime regulators within the water, energy and carbon cycles. If these functions are ignored, planners will be unable to assess, adapt to or mitigate the impacts of changing land cover and climate. Our call to action targets a reversal of paradigms, from a carbon-centric model to one that treats the hydrologic and climate-cooling effects of trees and forests as the first order of priority. For reasons of sustainability, carbon storage must remain a secondary, though valuable, by-product. The effects of tree cover on climate at local, regional and continental scales offer benefits that demand wider recognition. The forest- and tree-centered research insights we review and analyze provide a knowledge-base for improving plans, policies and actions. Our understanding of how trees and forests influence water, energy and carbon cycles has important implications, both for the structure of planning, management and governance institutions, as well as for how trees and forests might be used to improve sustainability, adaptation and mitigation efforts. [Ellison 2017: Abstract]

Ford 2017. Structural Complexity Enhancement (SCE) is part of a larger ecological concept: nature tends to complexity, providing its resiliency, flexibility and inventiveness. SCE in treatment of forests is a management approach that promotes development of late-successional structure, including elevated levels of coarse woody debris. It adds variety to tree ages (favoring older trees), and variations in available sunlight and habitat.

Large trees, previously assumed to slow in both productivity and growth rate (Weiner and Thomas 2001, Meinzer et al. 2011), function as long-term carbon sinks (Carey et al. 2001). These findings further support the significance of structural retention as a co-benefit to forest carbon storage. Adaptive silvicultural practices promoting multiple co-benefits, for instance, by integrating carbon with production of harvestable commodities, can contribute to efforts to dampen the intensity of future climate change while maintaining resilient ecosystems (Millar et al. 2007). Prescriptions that enhance in situ forest biomass and thus carbon storage offer one such alternative (Ducey et al. 2013). U.S. forests currently offset approximately 16% of the nation’s anthropogenic CO2 emissions, but this has the potential to decline as a result of land-use conversion and lack of management (EPA 2012, Joyce et al. 2014). While passive or low-intensity management options have been found to yield the greatest carbon storage benefit, assuming no inclusion of substitution effects (Nunery and Keeton 2010) or elevated disturbance risks (Hurteau et al. 2016), we suggest the consideration of SCE to enhance carbon storage. Multiple studies have explored co-benefits provided by management for or retention of elements of stand structural complexity, including residual large living and dead trees, horizontal variability, and downed CWM (Angers et al. 2005, Schwartz et al. 2005, Dyer et al. 2010, Gronewold et al. 2012, Chen et al. 2015). Silvicultural treatments can effectively integrate both carbon and late-successional biodiversity objectives through SCE based on this study and previous research (e.g., Dove and Keeton 2015). Remaining cognizant of the potential for old-growth compositional and structural baselines to shift over time and space with global change—climate impacts on forest growth and disturbance regimes, altered species ranges, and the effects of invasive species—will be important for adaptive management for late-successional functions such as carbon storage. [Ford 2017: 16]

Healing Harvest Forest Foundation. 

The spot compaction of animal feet is far less damaging to the forest soil and tree roots than the continuous track created by a wheel or track driven machine. Small sized tracts of timber can not be harvested with conventional methods that require higher capitalization and expensive moving cost.   The economic pressure in conventional forest harvesting operations influences most loggers to feel that they must cut all the trees to make their work cost effective.  This restricts the silvicultural prescriptions available for the management of the forest….Our method of selecting individual trees on a “worst first” basis and limiting removal to no more than 30% retains the forested condition and is indeed improvement forestry…. The holes created in the forest canopy are substantial enough for “shade intolerant” species to regenerate naturally from seedlings of the superior specimens that are left in a healthy “good growing” condition.  We believe that basically the repair of the forest from previous “high grading” is best accomplished through several successive “low grading” harvests. [Healing Harvest 1999]

Makarieva 2007.  The authors examine ecological and geophysical principles to explain how land far inland away from the ocean can remain moist, given that gravity continuously pulls surface and groundwater into the ocean over time.

All freshwater on land originates in the ocean from which it has evaporated, is carried on air flux, and precipitates over the land. Coastal regions benefit from this cycle by their proximity to the ocean, yet in the absence of natural forests in coastal regions precipitation weakens as distance from the ocean increases, leaving inland areas arid. The authors propose the concept of a biotic pump to explain how large continents can be sufficiently moist deep into the interior and abundant with rivers and lakes.

Air and moisture are pulled horizontally by evapotranspiration from coastal forests.  When water vapor from plants condenses, it creates a partial vacuum which pulls water evaporating from the ocean into the continental interior where it rains in forest.  By contrast, deserts are unable to pull ocean evaporation to them because they lack any evaporative force.

Such ongoing deforestation, and crucially coastal deforestation on a large scale, threatens to cut off rain to the interiors of Earth’s continents thereby creating new deserts. The Amazonian rainforest is the prime example.  Deforestation of the eastern coast of South America has led to changes in the rainforest that is resulting in drying and desertification of the interior, with unprecedented fires and loss of rivers.  Historically, Australia’s interior became a desert around the time the first humans arrived on the continent, and the authors speculate that early coastal deforestation was the cause. On the other hand, restoring natural coastal forests can also restore inland water cycles and reverse desertification.

This article illustrates the importance of biological relationships that are ecologically complex and poorly understood. It highlights the significance of the precautionary principle in assessing what we don’t know (and what we don’t know that we don’t know) when altering ecological processes, and taking preventive action in the face of uncertainty.