Mycorrhizal and Microbial Inoculation Affect the Growth and Survival of Native Plants Raised for Restoration

Item

Title

Eng Mycorrhizal and Microbial Inoculation Affect the Growth and Survival of Native Plants Raised for Restoration

Date

2014

Creator

Eng Porter, Sasha R

Subject

Eng Environmental Studies

extracted text

MYCORRHIZAL AND MICROBIAL INOCULATION
AFFECT THE GROWTH AND SURVIVAL OF NATIVE PLANTS
RAISED FOR RESTORATION

by
Sasha Porter

A Thesis
Submitted in partial fulfillment
of the requirements for the degree
Master of Environmental Studies
The Evergreen State College
December 2014

 

ABSTRACT
Mycorrhizal and Microbial Inoculation Affect the Growth and Survival
of Native Plants Raised for Restoration
Sasha R. Porter
Production of native seedlings for field outplanting has become a common ecological
restoration technique worldwide. However, the establishment of greenhouse-raised plants
in the field is usually poor. Mycorrhizal fungi are symbionts that can provide survival
benefits to host plants. This relationship is ubiquitous in terrestrial ecosystems and
mycorrhizae are absent only under unusual circumstances, such as in a nursery
greenhouse.
Nine plant species native to the highly endangered Northwest short-grass prairie and oak
savanna ecosystems (Balsamorhiza deltoidea, Castilleja levisecta, Dodecatheon
hendersonii, Dodecatheon pulchellum, Festuca roemeri, Gaillardia aristata, Micranthes
integrifolia, Ranunculus occidentalis, and Silene douglasii) were grown for six months in
sterilized medium with an arbuscular mycorrhizal fungi (AMF) inoculant cultured from
local native plants, a general AMF inoculant, or in control treatments. Three microbial
inoculants with AMF removed, created from a nearby site considered to be high-quality
remnant prairie, a restoration site, and unsterilized potting medium, were added within
each AMF treatment in a full factorial design. Seedling emergence, survival,
aboveground growth, and biomass data were collected, and remaining plants were
transferred to field sites for long-term monitoring.
AMF significantly enhanced the growth of five species and the survival of four, with no
detectable effect on the remainder. Further, there was no significant difference between
the two AMF inoculants. Field microbial wash tended to have a negative effect on
seedling emergence and growth, with the high-quality site treatment most repressive.
AMF and the introduced microrganisms interacted on Festuca roemeri, with AMF
mediating the negative effect of other fungi. Surprisingly, AMF positively affected the
growth of Castilleja levisecta, a hemiparasite, and altered the phenology of Dodecatheon
hendersonii, delaying dormancy. These results suggest that AMF can enhance the growth
and survivorship of many species, and that inoculation may lead to greater success in
ecosystem restoration efforts.


 

TABLE OF CONTENTS
FRONT MATTER . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ii
Abstract . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .ii
Table of Contents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .iv
List of Figures . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .vii
List of Tables . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .vii
Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .ix
Abbreviations and Definitions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .x
INTRODUCTION . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .1
I. LITERATURE REVIEW. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4
Native Plant Reintroduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .4
Mycorrhizal Fungi . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7
History and biology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7
Plant interaction with AMF . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10
AMF and other microorganisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16
Ecosystem effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17
Common Mycorrhizal Networks . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19
Nutrient transfer . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .21
AMF and Rare Plant Reintroduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24
Inoculant source . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .27
Risk . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .29
Pacific Northwest Prairie-Oak Ecosystems . . . . . . . . . . . . . . . . . . . . . . . . . . . . .30


 

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II. ARTICLE MANUSCRIPT: Mycorrhizal and Microbial Inoculation Affect the Growth
and Survival of Native Plants Raised for Restoration
Abstract . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .37
Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .37
Materials and Methods
Seed stratification and sowing . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 41
Growing medium and inoculants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 42
Experimental design . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 44
Data collection and analyses . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 45
Results
Growth . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 46
Emergence and survival . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 52
Discussion
AMF inoculation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .54
Soil microbial wash . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 56
Species-specific responses . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 57
Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 59
Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .61
Literature Cited . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .61
Supplementary Tables and Figures . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .66


 

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III. Restoration and its Discontents: A Critical Analysis of Ethics and Practice in Applied
Ecology
Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .68
Authenticity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .69
Humans, Nature, and Lines . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .70
Indigenous people . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .70
Work . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 71
Malicious Restoration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .72
Problems with Benevolant Restoration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .76
Invasive species and dates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 77
Herbicide dependency . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .79
Novel Ecosystems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .80
Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 82
THESIS REFERENCES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .85


 

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LIST OF FIGURES
LITERATURE REVIEW
Figure 1. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8
Microscopic view of a maize root (cleared) colonized by arbuscular mycorrhizal fungi (dyed blue)
(Hazel Davidson).

Figure 2. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11
Trends from a meta-analysis show effects of variables on plant response to arbuscular mycorrhizal
inoculation (Hoeksema et al. 2010).

Figure 3. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13
Sixty-four plants were inoculated with the arbuscular mycorrhizal fungi Glomus etunicatum and showed
high variability in growth response compared to non-inoculated plants (Klironomos 2003).

Figure 4. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22
Results of inoculation with a common mycorrhizal network (CMN) on growth of a C3 plant, flax (Linum
usitatissimum), and a C4 plant, sorghum (Sorghum bicolor) (Walder et al. 2014).

Figure 5. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 31
Geographical extent of the Willamette Valley–Puget Trough–Georgia Basin Ecoregion (Hamman et al.
2011).

MANUSCRIPT
Figure 1. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 47
Growth response of seedlings to inoculation with arbuscular mycorrhizal fungi at 16 weeks.
Figure 2. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 48
ANOVA showed that arbuscular mycorrhizal fungi (AMF) mediated the negative effects of microbial wash
on Festuca roemeri.

Figure 3. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 49
Roots of Micranthes integrifolia harvested at six months from plants A) in a nonmycorrhizal control and
B) inoculated with arbuscular mycorrhizal fungi

Figure 4. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 49
Belowground biomass of Micranthes integrifolia and Dodecatheon henderonii were significantly increased
by inoculation with arbuscular mycorrhizal fungi.

Figure 5. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 50
Correlation of root biomass and week entering dormancy for Dodecatheon hendersonii.

Figure 6. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 51
Growth response of seedlings at 16 weeks to inoculation with three microbial washes.

Supplemental Figure 1. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 66
Perennial Pacific Northwest Willamette Valley–Puget Trough–Georgia Basin species studied.


 

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Supplemental Figure 2. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .67
Treatments were arranged for each tray to contain one mycorrhizal treatment, one microbial treatment,
and three species.

LIST OF TABLES
Table 1. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .42
Ten native northwest prairie species were selected for this study.
Table 2. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .50
Results of two-way ANOVA were significant for interaction between arbuscular mycorrhizal fungi and
microbial source for Festuca roemeri.

Table 3. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 53
Chi-square analysis of the effects of arbuscular mycorrhizal fungi (AMF) inoculation and microbial wash
on a) seedling emergence and b) survival.

Supplemental Table 1. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 66
Seeds underwent imbibition (soaking) and stratification (dark storage at 3° C) based on previously
established protocols.


 

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ACKNOWLEDGEMENTS
This work would not have been possible without the support of many organizations and
amazing people. It was an absolute pleasure to work with Dr. Erin Martin and Dr. Sarah
Hamman, both of whom contributed greatly to the ideas, methods, and writing of this
thesis. The Center for Natural Lands Management (CNLM) and the Evergreen Science
Support Center were also instrumental. I would like to specifically acknowledge Sierra
Smith, Tel Vaughn, Carl Elliot, Scott Morgan, Hansina Hill, Kaile Adney, and CNLM
interns. My partner, family, and friends provided amazing support throughout the fifteen
months I spent working on this project and deserve more thanks than I can acknowledge
here.
I am grateful for the financial assistance supplied by the Washington State
Department of Agriculture Nursery Research Grant and a fellowship from the Evergreen
State College Office of Sustainability.
Finally, I would like to acknowledge the indigenous people of the Pacific
Northwest who lived in and maintained the prairie and oak savannas for thousands of
years, and whose people and culture, like their lands, have been unquantifiably damaged
by invasion that continues today.


 

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ABBREVIATIONS AND DEFINITIONS
AMF: arbuscular mycorrhizal fungi. A phylogenetic group (Glomeromycota) of fungi
composed of branching hyphae that enter the roots of plants and form a symbiotic
relationship in which water and nutrients are exchanged with autotrophs for carbon
AQFO: Aquilegia formosa Fisch. ex DC. “Western columbine”
Autotroph: Plant that photosynthesize
BADE: Balsamorhiza deltoidea Nutt. “Deltoid balsamroot”
CALE: Castilleja levisecta Greenm. “Golden paintbrush.” An Endangered Species Act–
listed hemiparasite that can photosynthesize but also gains nutrition by extending rootlike
organs called haustoria into the belowground systems (but not cells) of other plants.
CNLM: The Center for Natural Lands Management.
DOHE: Dodecatheon hendersonii A. Gray. “Mosquito bills”
DOPU: Dodecatheon pulchellum (Raf.) Merr. “Darkthroat shootingstar”
GAAR: Gaillardia aristata Pursh. “Blanketflower”
Hemiparasite: a plant that can photosynthesize but often gains nutrition through feeding
off the roots of other plants
Inoculant: mycorrhizal and/or microbial cultures added to soil growing medium.
JBLM: Joint Base Lewis McChord. The owner of the largest intact remnants of Pacific
Northwest prairie. This large site south of Tacoma was acquired for Department of
Defense use early in the 20th century due to its open landscape, and much of the area has
ironically been maintained by the constant setting of fires from artillery since.
MH: mycoheterotrophic plant. A plant that does not photosynthesize for itself and
instead gains nutrition by accessing carbon allocated to mycorrhizal fungi by other plants
through the exploitation of mycorrhizal networks
MIIN: Micranthes integrifolia (Hook.) Small. “Wholeleaf saxifrage”
mycorrhiza pl. mycorrhizae: literally “fungal root.” Several functional groups of
mycorrhizal fungi exist, all form symbiotic relationships with plants
MN: mycorrhizal network. A condition under which multiple plants are linked by a one
mycorrhizal fungus through its hyphae.
NM: non-mycorrhizal. An unusual state in which a plant cannot or does not form a
relationship with mycorrhizal fungi.
RAOC: Ranunculus occidentalis Nutt. “Western buttercup”
SIDO: Silene douglasii Hook. “Douglas’s catchfly”
WPG: Willamette Valley–Puget Trough–Georgia Basin. An ecoregion containing
distinct and threatened prairies and oak savannas


 

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INTRODUCTION
This thesis explores the possibility of using microscopic organisms to help solve some
very large problems. Globalization and human exploitation of fossil fuels are causing
constant and unpredictable environmental crises on earth. Among these, the loss of
important natural and cultural landscapes, and a steady decline in biodiversity globally,
are especially concerning, and many scientists suggest that we are in the midst of Earth’s
sixth great extinction event (Barnosky et al. 2011). Current human practices are not
predicted to experience major positive changes in the near future and the effects of
increasing CO2 will exacerbate environmental conversion and the loss of biodiversity
(Thomas et al. 2003). In the face of this dire future conservationists are working hard to
mitigate changes by restoring habitats and attempting to prevent extinction.
Encouraging the re-establishment of viable wild populations of rare and native
plant species through cultivation and outplanting to historical habitats is a widespread
restoration technique (Machinski & Haskins 2012). Plant reintroduction attempts to
mitigate loss of biodiversity and to prevent extinction by increasing native plant
abundance and diversity, with a resultant preservation of species across trophic levels.
The propagation of native plants for restoration is a prevalant, accepted practice that is
generally unsuccessful, with very few reintroduced plants surviving more than a year or
two, and even fewer establishing, flowering, or fruiting over time (Godefroid et al. 2010).
Failure may be due to horticultural techniques that provide abundant nutrients but poorly
emulate native environments (Haskins & Pence, 2012).
This manuscript-style thesis investigates whether inoculating growth medium
with mycorrhizal fungi and other rhizosphere microorganisms may provide plants with

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important traits, adaptations, and tools that will lead to greater long-term establishment in
the field. The first chapter, a literature review, begins by exploring some of the problems
that the practice of native plant reintroduction has experienced. It provides background
on the evolution and biology of mycorrhizal symbiosis, a relationship between autotrophs
and mutualistic fungi in which most terrestrial plants engage, and a more detailed
analysis of the effects of arbuscular mycorrhizal fungi (AMF) on plants and ecosystems.
Previous AMF inoculation research, including the formation of mycorrhizal networks
(MN) between multiple plants, and emerging themes from the literature are included.
Few published studies on the use of AMF for rare and native plant propagation
exist, but the compelling findings and problems of several of these are explored, as are
potential AMF sources and associated costs and risks. AMF engage in complex
relationships with other microorganisms, and literature related to rhizosphere interactions
and their application to research methods are reviewed. Finally, the highly endangered
prairie-oak savannas of the Pacific Northwest, USA, and the potential use of mycorrhizal
inoculation techniques as part of the restoration strategy for this rare ecosystem is
proposed.
The second chapter, an original research manuscript formatted for publication
presents the results of an experiment that addresses the following questions 1) How does
AMF inoculation affect the growth of greenhouse-raised seedlings? 2) Does AMF affect
short-term (6 months) survival? 3) Is an AMF inoculant cultured from native soils
superior to a commercially available one? 4) How will AMF inoculants interact with
different soil microbial communities likely to be present in outplanting sites?

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While the results found in the manuscript raise interesting basic science questions,
the work was done specifically to provide information for practice. Many of the ideas in
this thesis are rooted in the field of restoration ecology, and the third chapter analyzes
some of the ethical and practical issues associated with this emerging science. This
chapter integrates the thoughts of both philosophers and scientists to explore the changing
relationship between humans and nature, important criticisms to the ideas at the
foundation of restoration, and some of the problems with ecological restoration as
currently practiced.
Mycorrhizal symbiosis and ecological restoration are highly complex processes
that need to be understood within the context of theory. The complexities inherent to both
this keystone biotic relationship, and restoration ecology itself, will undoubtedly be
exacerbated by human-environmental conflict and a changing climate. A growing
understanding of environmental processes, including plant-microorganism interactions
and the possibilities and limitations of applied ecology, will be useful to creating positive
change in the future. The native plants grown as part of this thesis have been transferred
to field sites for long-term monitoring, and I hope that both these seedlings, and the ideas
of this work will continue to thrive and have a positive effect on our world.

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Literature Review
NATIVE PLANT REINTRODUCTION
Reintroduction of rare, native, and endangered plant species has become an important
restoration tool worldwide. Success can prevent extinction, benefit species across trophic
levels, and restore ecosystem functionality to degraded sites (Maschinski & Haskins
2012). The actual success rate of plant reintroductions, however, is likely quite low
(Godefroid et al. 2010). Long-term monitoring of the outcome of reintroduction efforts is
infrequent, and the published literature reflects a strong bias toward positive results
(Godefroid et al. 2010; Drayton & Primack 2012). In a meta-analysis, Godefroid et al.
(2010) compared data from twenty-six published papers with results from a survey sent
to 473 institutions that were suspected of having participated in reintroductions without
publishing results. It was found that survival rates in the literature were much higher
(78%) than those reported in survey data (33%), and that in studies where longer-term
results were available, a startling decline in success occurred over time with an average
of only 6% of reintroduced plants flowering after 4 years (Godefroid et al. 2010).
A variety of suspected and unknown factors affect the success of rare plant
restoration efforts and the science itself is still young (Dalrymple et al. 2012). There is a
great need for the development of techniques that will increase the viability of
reintroduction efforts and research should occur at the species, ecotone, and global scales
(Godefroid et al. 2010; Dalrymple et al. 2012). Greater understanding of the effectiveness
and possible repurcussions of restoration methods at a variety of scales, and the
variability within and between systems may allow practitioners to more easily and
effectively implement successful projects, even where studies have not been conducted.

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Restorationists working with rare and endangered species are often under pressure to
produce short-term results due to the imminent threat of extinction. This can lead to
haphazard restoration efforts that do not provide useful empirical data to guide future
projects. Guerrant (2012) argues that regardless of outcome, reintroduction efforts need
to be structured as designed scientific experiments in order to produce reliable and
replicable results.
Long-term monitoring of reintroduction efforts and the publication of results is
important as “failures” can often provide more-valuable information for the development
of technical strategies than successes (Drayton & Primack 2012). In an unusual case
where a reintroduction effort was recensused after 15 years, Drayton and Primack (2012)
were surprised to discover that populations that were considered well-established three
years after planting had almost entirely disappeared after 15 years. In the study,
wildflower species planted in 1995 were resurveyed after two years and as seven of the
eight species were present at reintroduction sites, leading the authors to publish the
results as successful (Drayton & Primack 2000). In 2010, however, no surviving
individuals of six of the eight species were found, and a seventh species was present at
only one site, leading the researchers to conclude that long-term success rates for
establishing new plant populations are very low, even when efforts are initially
considered successful, and that research into factors that affect establishment over time is
urgently needed (Drayton & Primack 2012).
Horticultural techniques that emphasize growth rates and short-term survival (one
to two years) are often used in the cultivation of seedlings for restoration efforts. Seed
stock, time, and funding are regularly in short supply, and propagating plants in

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greenhouses with abundant nutrients, water, and lack of competition produces the largest
number of seedlings for outplanting. However, high rates of failure occur when these
propagules are introduced to stressful field conditions, which likely involve competition
with other species, dense and low-nutrient soils, herbivory, drought, and pathogens
(Godefroid et al. 2010; Haskins & Pence 2012). Restoration sites are by definition altered
versions of the environments that plants evolved in, with additional anthropogenic
stresses that often include invasive species, and changes to natural disturbance regimes,
with soils affected by both these alterations and the restoration process itself.
Propagation methods that emulate field conditions can reduce the stress of
acclimatization by providing plants with tools such as adaptive root structures and
symbiotic partners that allow increased access to nutrients (Haskins & Pence, 2012).
Nutrient stress may be especially problematic in restoration sites with a history of
invasion. Non-native species often alter soil chemistry, for example, nitrogen-fixers have
been shown to decrease soil phosphorus (Thorpe et al. 2013). Inoculation with
mycorrhizal fungi often has positive benefits for plant survival, but the majority of
research on mycorrhizal relationships has occurred in agriculture, horticulture, forestry,
and basic science rather than conservation or restoration (Haskins & Pence 2012). The
use of mycorrhizal fungi in the propagation of rare plants shows promise and may be
especially suited to solving problems with acclimatization and long-term survival in
reintroduction efforts (Gemma et al. 2002; Rowe 2007; Zubek 2008; Ferrazzano &
Williamson 2013).

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MYCORRHIZAL FUNGI
History and biology
Mycorrhizal fungi form symbiotic relationships with 80–90% of land plants and exist in
nearly every terrestrial ecosystem (Smith & Read 2008). It is not through roots, but
extensive mycorrhizal hyphal networks that plants uptake the majority of necessary
nutrients, with autotrophs donating carbon to fungi in exchange for resources (Smith &
Read 2008). Paleobotanical, morphological, and DNA-based evidence indicate that the
earliest plants formed relationships with arbuscular mycorrhizal–like endophytes 400
million years ago, long before the evolution of roots (Brundrett 2002). Fungi likely
originated over one billion years ago, predating terrestrial colonization (Smith & Read
2008). This ancient symbiosis heavily influences nutrient cycling, plant community
structure, diversity, and soil characteristics today, and is considered to have a keystone
ecological function globally (van der Heijden 1998; Jeffries et al. 2003).
Smith and Read classify mycorrhizal fungi into seven functional groups based on
structural characteristics and autotrophic associates. Among these groups, the arbuscular
mycorrhizal fungi (AMF) are by far the most abundant, and were recently organized into
a separate fungal phylum, Glomeromycota, based on DNA sequencing (Schüβler et al.
2001). Unlike other mycorrhizal fungi, which are specific in host selection and only able
to form partnerships with certain types or families of plants, AMF are generalists and
capable of forming biotrophic relationships with an extremely wide range of autotrophs
(Smith & Read 2008). AMF occur in almost all vegetated terrestrial areas and are the
dominant mycorrhizal type in grasslands, tropical forests, and agricultural systems (Smith
& Read 2008).

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Morphologically, AMF consist of mycelium formed by masses of branching,
threadlike hyphae that extend both into soil and within the roots of plants (Smith & Read
2008). Unlike ectomycorrhizal fungi, which form sheaths around root tips, AMF extend
hyphae into and between the cortical cells of roots, forming intracellular arbuscules
(branching tree-shaped organs) through which resources are passed bidirectionaly (Fig. 1;
Smith & Read 2008).

Figure 1. Microscopic view of a maize root (cleared) colonized by arbuscular mycorrhizal fungi (dyed
blue). Structures within the root including vesicles, arbuscules, and hyphae, extraradical hyphae are
also visible. Image adapted from Hazel Davidson, University of Aberdeen.

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Unlike some saprotrophic fungi and ectomycorrhizae, under most circumstances
AMF mycelia cannot be seen with the naked eye and do not form epigeous sporocarps of
fruiting bodies such as mushrooms, but reproduce instead through large thick-walled
spores spread by hyphae and fauna (Smith & Read 2008). Despite being very small (2–20
µm) AMF hyphae are extensive and ubiquitous, contain recalcitrant compounds, form
unique conglomerates, live only 5–7 days on average, and thus likely contribute large
quantities of organic carbon to soils (Staddon 2003; Smith & Smith 2011).
Researchers have only begun to tease out the actual genetic, cellular, and
molecular interactions that allow the formation of mycorrhizae in the last decade or two,
and details on what has been found could fill many much longer literature reviews (Smith
& Read 2008). However, it is known that both AMF and plants respond to each other
through a variety of complex signaling pathways and gene expression, which act in a
coordinated manner to form mycorrhizae (Smith & Read 2008). Preformation signaling
occurs, but precise mechanisms are not yet entirely understood, and this signaling is also
stimulated (or potentially suppressed), by other microorganisms, such as the
mycorrhization helper bacterium AcH 505 (Kurth et al. 2013). Interestingly, parasitic
plants may have evolved to outilize the same pathways that mediate AMF recognition
and colonization in host plants in order to exploit neighboring autotrophs as a carbon
source without reciprocation (Fernández-Aparicio et al. 2010).
Mycorrhizal fungi are heterotrophic and rely on organic carbon from their
photosynthetic partners (Smith & Read 2008). The entirety of resources that plants
receive from this symbiosis, in contrast, is complex, and to what degree it is beneficial at
any given time is not yet fully understood (Smith & Smith 2011). Water, phosphorus (P),

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nitrogen (N), and trace minerals including copper (Cu) and zinc (Zn) are also involved
(Smith & Read 2008). Isotopes have been used effectively in elucidating some of the
details of nutrient transfer, but larger effects on plants such as growth, survival, diversity,
hormone levels, architecture, tolerance to toxins, and disease- and drought-resistance also
occur, but are less fully understood (Hartnett & Wilson 2002; van der Heijdan 2004;
Smith & Smith 2012).
Reductionism has its place in understanding the benefits of mycorrhizal
association to plants, but complexity theory and its tenets, including nonlinearity, positive
and negative feedbacks, network connections, and emergent traits are extremely useful in
understanding these relationships and their context within a larger ecological framework.
The details of the transfer of benefits between symbiotic partners is an inherently
complex relationship in a system that involves a variety of interacting dynamics. These
are known to include plant variety and condition, the likely presence of many differently
acting mycorrhizal species, changing resource levels, plant age and community,
ecological conditions, the presence of other microorganisms and pathogens, and
feedbacks between these and other unknown elements (Smith & Smith 2012; Hartnett &
Wilson, 2002).

Plant interaction with AMF
In a systematic meta-analysis of plant response to mycorrhizal inoculation from 1,994
studies in 183 publications, Hoeksema et al. (2010) found a large variation in results with
certain patterns emerging (Fig. 2). Plant functional group was the most important
explanatory variable in AMF experiments, with (non-N-fixing) forbs exhibiting the

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Figure 2. Trends from a meta-analysis show effects of other variables on plant response to
arbuscular mycorrhizal inoculation. Figure from Hoeksema et al. (2010).


 
highest level of positive response to mycorrhizal association (Hoeksema et al. 2010).
Whether plants were inoculated with one or more species of AMF was also important
with inoculants containing more species promoting the greatest plant response, defined in
this study as “the log response ratio of inoculated to non-inoculated plant biomass”
(Hoeksema et al. 2010). The presence of other rhizosphere microorganisms significantly
positively affected plant growth with AMF compared to sterile soil and shows that
important and complex interactions occur between AMF, plants, and other members of
the microbial community (Hoeksema 2010; Philippot 2013). Kurth et al. (2013) showed
that a bacteria, AcH 505, is fungus-specific and produces growth regulators that
significantly stimulate mycorrhizal formation when roots are inoculated with the microbe,
which also leads to an increase in plant growth. Further, mycorrhization helper bacterium

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have exhibited both a longer lifespan and increased abundance with greater MF presence
(Kurth et al. 2013).
Hoeksema et al. (2010) found that N fertilization of soil significantly influenced
plant response to AMF with growth showing a more positive response to fungal
inoculation without fertilization with N. Where N is abundant, mycorrhizal relationships
are reduced, likely because the symbiosis is less necessary for N acquisition. The
introduction of N into ecosystems through either nitrogen-fixing invasive plants or
agriculture can lead to a reduction in mycorrhizal fungi throughout the system
(Vogelsang & Bever 2009, Thorpe et al. 2013). The majority of research on suppression
of mycorrhizal fungi by N addition has occurred in agriculture, where the addition of Nbased fertilizers suppresses AMF and changes systems from fungally to bacterially
dominated, and requires the addition of greater amounts of P as non-mycorrhizal plants
are less able to access this nutrient (Six et al. 2006). These changes in turn negatively
affect soil aggregation, C storage, and nutrient leaching; these effects may increasingly
spread from agricultural to wild systems with global changes to N cycling (Six et al.
2006; van der Heijden 2010; Asghari & Cavagnara 2012). AMF suppression through N
fertilization is likely also problematic for plants grown in pots and then outplanted into
environments that may have increased N from invasive plants, an agricultural history, or
proximity to agriculture.
Plants show a high level of variability in response to AMF based not just on
functional type, but individual plant species. AMF species can also affect response
(Smith & Read 2008). Klironomos (2003) tested 64 plant species with an inoculant of
one AMF species, Glomus entunicatum, and found plant growth responses that varied

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from highly positive to highly negative (Figure 3). While the majority of responses were
not significantly different from a non-inoculated control, a clear pattern of variability
emerged when comparing among species. A negative growth response to AMF is always
a possibility, and can occur when plants donate carbon for non-limiting nutrients.
Wilson and Hartnett (1998) found similar variability testing 95 tallgrass-prairie
species in a greenhouse study. When categorized by plant functional type, forbs and C4

Figure 3. Sixty-four plants were inoculated with the arbuscular mycorrhizal fungi Glomus
etunicatum and showed high variability in growth response compared to non-inoculated plants.
While results from individual plants were not necessarily significant, the data show a clear trend of
variability. Figure from Klironomos (2003)

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grasses benefited most from AMF association, while C3 grasses tended to have a neutral
response and legumes (N-fixers) had a significant negative response (Wilson & Hartnett
1998). The high level of mycorrhizal obligation found in C4, as opposed to C3, grasses
may be related to coarser root structures with a resultant decrease in nutrient acquisition
ability, most prolific growth during summer, rather than spring, when water may be a
limiting factor, and an abundance of C to allocate due to the more efficient
photosynthetic pathway utilized by these plants.
Interestingly, native tall-grass prairie perennials were significantly positively
affected by AMF while annuals showed lower root colonization and were not positively
affected (Wilson & Hartnett 1998). The authors theorize that this is because annuals tend
to thrive in newly or regularly disturbed areas where there may be less competition, and
an advantage to growing very quickly without the need for long-term survival strategies.
There also tends to be less AMF presence immediately after some disturbances, so
annuals may have developed evolutionary strategies that were successful without AMF
present, and that in fact gave a short-term advantage over late-successional species
(Jasper 1991).
Response to AMF is complicated not only by plant type, but by interactions
between plants and AMF species. In another experiment, Klironomos (2003) grew 10
plants with each of 10 AMF species and found that for a single plant response could vary
from highly positive to highly negative depending on the mycorrhizal fungi species it was
inoculated with. Plantago lanceolata, for example, showed a negative response of ~45%
when inoculated with Acaulospora morrowiae and a positive response of ~45% when
inoculated with Glomus mosseae (Klironomos 2003). No plant or AMF species

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consistently responded either positively or negatively, and in fact all 10 species showed
both positive and negative reactions to different AMF species, though not all of these
results were statistically significant (Klironomos 2003). Interestingly, even genetically
different individuals of the same AMF species affected plants in unpredictable ways and
could cause a positive or negative response. Responses were most extreme with AMF
genets from the site the seeds were collected from, as opposed to a foreign site
(Klironomos 2003). The author concludes that variability in plant response to different
AMF species may functionally maintain plant diversity on sites with multiple AMF
species, but the exact mechanisms causing the variability are not well understood
(Klironomos 2003).
Genetic variation within plant species can also affect response to AMF (Anderson
& Roberts 1993). In a study of the prairie grass Schizachyrium scoparium, seeds from
three locations were inoculated with AMF cultured from one of the source areas.
Seedlings from the AMF-origin site and a nearby site (both in Mason County, IL) had
significantly greater biomass than seedlings from a Nebraska source, and all three were
larger than controls grown in sterilized soil (Anderson & Roberts 1993).
A theoretical model in which plant-mycorrhizal relationships are viewed as
existing on a continuum between parasitism and mutualism has become widespread in the
literature (Johnson et al. 1997). This model is useful in understanding variation in plant
responses to mycorrhizal fungi, and rare cases where mycorrhizae formation causes clear
negative effects, but the authors acknowledge that defining costs and benefits, especially
at ecologically meaningful scales is difficult if not impossible (Johnson et al. 1997). The
continuum approach is likely a simplification of a nonlinear relationship and not useful in

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describing and predicting systems with high levels of complexity. The response of an
individual plant to AMF can change throughout its lifetime, fluctuating from a positive to
a neutral or negative response based on age, identity of the AMF, presence of pathogens
or herbivory, and stressors such as drought (van der Heijden 2004; Smith & Read 2008).
It is also a great simplification to define responses as positive or negative based only on
the commonly used factor of growth rate, or economic models of nutrient exchange,
when other morphological, qualitative, and unknown factors are also affected by
mycorrhizal fungi (Smith & Read 2008; Hartnett & Wilson 2002). Smith and Smith
(2012) also argue that even when there is a negative growth response, AMF are never
truly parasitic because there appears to always be P transfer from AMF to plant, while
parasitism implies unidirectionality of resources.

AMF and other microorganisms
Within a few centimeters of the roots of plants exists one of the most diverse and
dynamic systems on Earth (Philippot 2013). In this highly complex interface, known as
the rhizosphere, plants, fungi, bacteria, and other organisms interact with one another
with effects that regulate the growth, composition, and biomass of plants, which directly
or indirectly affect other organisms, making this one of the most important systems on
the planet. Almost all organic nitrogen is first fixed by bacteria and archaea in the
rhizoshpere, and carbon delivered by photosynthesizers enters this zone through roots and
AMF (Chapin 2011). The intensified biogeochemical cycling of the rhizosphere
combined with the multitude and variability of species and relationships that may be
present make it a difficult area to study.

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When conducting mycorrhizal research scientists often sterilize some portion of
soil or growing medium to establish non-mycorrhizal control(s) (Koide & Li 1989). This
methodology, however, also eliminates non-mycorrhizal organisms from control groups
leading to unbalanced experiments if one treatment contains a whole soil community
while another contains either no microorganisms or only added AMF without other
microorganisms that would occur in natural soil. This problem can be exascerbated when
whole-soil is sterilized, as elements, especially manganese, which can be toxic at high
levels, are released by the autoclaving process (Koide & Li 1989).
Scientists have dealt with this issue in a variety of ways; one common method is
to sterilize potting soil, and then add mycorrhizal inoculant to treatment groups, and a
“soil microbial wash” to all groups. This wash is created by making a soil slurry with
water, and filtering it through a sieve that is 38µm or less, which removes AMF spores
but allows many other microorganisms to pass through (Koide & Li 1989). The resulting
wash may contain beneficial microorganisms, such as mycorrhizal helper bacterium or
pathogenic fungi such as rusts (Kurth et al. 2013). Microbial washes tend to have a
beneficial effect on plant growth when combined with AMF (Hoeksema et al. 2010).

Ecosystem effects
AMF also have important effects at the ecosystem-scale (Rillig 2004; van der Heijden
1998; Wagg et al. 2011). In a study where 11 plant species were combined to simulate
European calcareous grasslands, it was found that AMF species diversity significantly
positively influenced plant diversity and ecosystem functioning (van der Heijden et al.
1998). Microcosms were inoculated with one of four native AMF species, a combination

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of the four, or a nonmycorrhizal control, and it was found that low AMF diversity led to a
few species becoming dominant at the expense of others (van der Heijden et al. 1998). A
parallel macrocosm study simulating North American old-field ecosystems found that
both plant biodiversity and ecosystem richness increased with the number of AMF
inoculant species (van der Heijden et al. 1998). These combined results led the
researchers to suggest that AMF should be considered as determinants of plant diversity
in natural ecosystems (van der Heijden et al. 1998).
Hartnett and Wilson (1999) also found a strong effect of AMF on plant
community diversity in a five-year tallgrass-prairie field study; however, this ecosystem
experienced a significant decrease in plant biodiversity with mycorrhizae. Suppression of
AMF by fungicide in the field resulted in a large increase in plant species biodiversity,
and no change in aboveground biomass (Hartnett & Wilson 1999). The authors theorize
that the elimination of AMF led to a decrease in obligately mycotrophic C4 tall grasses,
with subsequent increases in subordinate C3 grasses and forbs (Hartnett & Wilson 1999).
A secondary finding indicates that across the five-year study period, annual precipitation
was negatively associated with mycorrhizae in that AMF root colonization increased with
decreasing annual precipitation. While the mechanism behind an increase in colonization
was likely because plants were better able to access water, a limiting resource, through
more extensive AMF networks, this also further elucidates the complexity of mycorrhizal
symbiosis (Hartnett & Wilson 1999).
More recently, Vogelsang et al. (2006) found that plant diversity and productivity
were more responsive to AMF identity rather than diversity. In addition, the authors
found that complex interactions occur among AMF species and P sources that alter

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community-level ecosystem properties. Increasing types of P source (from one to five,
both organic and inorganic) added to the ability of AMF to promote diversity in plants
(Vogelsang et al. 2006). AMF presence and diversity was found to reduce plant–plant
competition in a greenhouse co-planting study (Wagg et al. 2011). Four AMF species
were tested individually and in combination to elucidate growth dynamics between two
plants. AMF diversity was shown to reduce competition by reducing the growth
suppression effects of a grass on a legume in soils of varying quality, and the authors
suggest that a species-rich AMF community may act as insurance in maintaining plant
productivity in a fluctuating environment (Wagg et al. 2011).
The body of work on the ecosystem effects of mycorrhizal fungi is young,
evolving, and undergoing a rapid increase, and sweeping conclusions cannot be drawn,
yet it is clear that AMF play a significant role in shaping and maintaining plant
communities. There appear to be complex feedbacks regulating these relationships, and it
has also been shown that plant community composition can affect AMF community
diversity (Eom 2000; Hausmann & Hawkes 2009). Greater AMF diversity seems to
create emergent and self-organizational effects, regulating ecosystems in a wide variety
of ways, as well as increasing growth effects on individual plants (van der Heijden et al.
1998; Hartnett & Wilson 1999; Wagg 2011; Hoeksema 2010).

COMMON MYCORRHIZAL NETWORKS
A single AMF may colonize more than one plant, creating a common mycorrhizal
network (CMN) and greatly increasing the complexity of symbiotic dynamics (Selosse
2006). Research indicates that plants share nutrients, water, and signals through CMNs

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(Egerton-Warburton et al. 2007; Song et al. 2010). In an experiment with tomato plants
linked only by a common mycorrhizal network, Song et al. (2010) found that when one
plant was exposed to a pathogen the neighboring plant, which had contact only through
mycorrhizae, released at least six enzymes and induced six genes related to defense.
Collectively plants are known to produce a staggering variety of root exudates, perhaps in
excess of 100,000 (Bais et al. 2004). Research by Song et al. (2010) and others shows
that plants are able to exchange these signals through CMN, likely incurring useful
warnings and benefits in a form of “communication” unavailable to plants that are not
part of CMNs (Simard et al. 2012). This sharing of signals may be one reason plants
associate with AMF even when it does not appear to be beneficial based on growth rate
or nutrient economics.
Much of the initial evidence and proof of the existence of CMNs comes from
studies of mycoheterotrophic (MH) plants (Simard et al. 2012; Courty et al. 2011). MH
plants live in the forest understory and obtain C through the exploitation of CMNs
maintained by photosynthesizers; this C-obtaining strategy has evolved independently
several times and some plants are also known to be partially MH, heavily supplementing
photosynthesis with resources from a CMN (Courty et al. 2011). In a recent study of MH
plants in the tropics, Courty et al. (2011) analyzed the stable isotopes 15N and 13C to
examine the nutritional dynamics of MNs. AMF have been shown to be depleted in the
stable isotope 13C by ~ -2‰ to -4‰ compared to their host plants, and MH plants reflect
the δ13C signatures of CMN rather than nearby plants (Etcheverria et al. 2009; Courty et
al. 2011; Walder et al. 2013). This change in 13C indicates that carbon is being transferred
from the host plant to a MN and subsequently to the MH plant. These isotope studies

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effectively show that plants are able to indirectly obtain resources from other plants via
CMN, and results from lab and field studies using autotrophs linked by CMN have shown
bidirectional resource transfer (Simard et al. 1997, 2012; Philip et al. 2010).

Nutrient transfer
In a series of experiments in the late 1990s a British research group analyzed C transfer
between autotrophs linked by a CMN through the analysis of δ13C signatures (Fitter et al.
1998; Watkins et al. 1996; Graves et al. 1997). Plants that use a C4 photosynthetic
pathway have distinctly more enriched δ13C signatures (Cynodon dactylon ~ -14‰) than
C3 plants (Plantago lanceolata ≈ -28‰) and this difference has been effectively used to
trace the C transferred through CMNs (Sage & Monson 1999). C4 photosynthesis has
emerged several times evolutionarily in different parts of the world, is considered more
efficient than C3 photosynthesis, and occurs in the majority of grasses, as well as crops
such as maize, sugarcane, millet, and sorghum (Sage & Monson 1999). Co-planting
experiments using C3 and C4 plants linked by a CMN took advantage of the natural
difference in δ13C signatures to effectively show that autotrophs transfer C to other plants
through CMNs (Fitter et al. 1998; Watkins et al. 1996; Graves et al. 1997). Whether or
not the majority of C transferred to plants via CMN stays in roots or enters shoots is still
debated among scientists, and it is clear that at different life stages plants may benefit
from C exchange more than at others (Simard et al. 1997, 2012; Philip et al. 2010).
A recent greenhouse experiment by Walder et al. (2012) elegantly showed that
plants do not contribute to and receive benefit from CMNs equally. A C3 plant, Linum
usitatissimum, and a C4 plant, Sorghum bicolor, were coplanted with mycorrhizae and in

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Figure 3. Results of inoculation with a common mycorrhizal network (CMN) on growth of a C3 plant,
flax (Linum usitatissimum), and a C4 plant, sorghum (Sorghum bicolor). Flax benefits positively from
both the CMN, and the relationship with sorghum, while sorghum experiences only small changes in
growth (Walder 2012).


 
monoculture with and without AMF (Walder et al. 2012). Hyphae from the CMN were
exposed to the stable isotopes 15N and 33P as tracers in a separate hyphal compartment
divided from the roots by a 25µm mesh screen that AMF hyphae, but not plant roots,
could cross (Walder et al. 2012). It was found that with inoculation of Glomus
intraradices the C4 plant donated 90% of the C to the CMN and received only 10% of the
N and P exchanged (Walder et al. 2012). Inversely, the C3 plant donated only 10% of the
total C to the CMN yet received 90% of the N and P from the network (Fig. 3; Walder et
al. 2012). This unequal relationship resulted in a 298% positive change in growth for L.

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usitatissimum, the C3 plant (Walder et al. 2012). A similar pattern, though less extreme in
disparity occurred with inoculation of a different AMF species, Glomus mosseae (Walder
et al. 2012). These results show clearly that autotrophs can exploit CMNs through
unequal contribution to and benefit from networks, regardless of whether C is actually
significantly assimilated into shoot matter, and contradicts assumptions that “fair trade”
exist in CMN symbiosis. Walder et al. (2012) theorize that C4 plants have a carbon
excess, and thus are trading in “luxury goods,” which are not actually needed, but that
these results may differ over the whole lifecycle of plants. The C4 species involved in the
experiment actually lost very little in terms of biomass compared to the monoculture, but
greatly improved the growth of the C3 plant, showing again that relationships are
nonlinear, involve feedbacks, and lead to facilitation and emergent qualities, such as the
298% increase in growth of the C3 plant with only a 7% reduction in the growth of the C4
species (Walder et al. 2012).
While it is still unclear how CMNs affect plants in the field, these networks may
be key to some relationships with bidirectional transfer affecting species composition. In
Canada, it was found that Douglas fir (pseudotsuga menziesii) and paper birch (Betula
papyrifera) exchange carbon through ectomycorrhizal fungi in a season-specific pattern
that benefits both species (Philip 2010). When P. menziesii was shaded in the summer, B.
papyrifera transferred more carbon to the conifer, and P. menziesii reciprocated by
donating C to the deciduous tree in fall and spring (Philip 2010). Bidirectional transfer of
resources may help to maintain diversity and stability in ecosystems, and re-establishing
these relationships through the reintroduction of mycorrhizal fungi could be key to
successful restoration efforts.

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AMF AND RARE PLANT REINTRODUCTION
Smith and Read (2008) argue that mycorrhizal colonization is normal for plants, and that
existing in a non-mycorrhizal state (NM) should be considered abnormal. The NM
condition occurs only under special circumstances, such as when a plant is one of
approximately 10% of species that are considered NM, where extreme disturbance to soil
has occurred, for example in sites degraded by mining, or when plants are grown in pots
and not inoculated with mycorrhizae (Smith and Read 2008). Therefore, plants grown in
greenhouses for reintroduction without mycorrhizae should generally be considered to be
developing under abnormal conditions and likely have differences in architecture, growth
rate, and pathogen resistance compared to mycorrhizal counterparts (Haskins & Pence
2012).
Several researchers have investigated using AMF inoculant in the propagation of
native, rare, and endangered plants in the greenhouse for eventual reintroduction, with
successful results (Haskins & Pence 2012). In Arizona, Richter and Stutz (2002)
inoculated Sporobolus wrightii, a formerly dominant grass species in semi-arid riparian
floodplains, with local AMF in a greenhouse experiment. Previously S. wrightii had been
directly seeded with virtually no success, and only mixed success had been seen for
propagation and transplanting. Seedling emergence in the greenhouse was higher in
inoculated pots, and more tillers were produced in inoculated plants grown in small pots,
though growth was not affected (Richter & Stutz 2002). After transplant to the field, S.
wrightii seedlings propagated with AMF showed greater survival, basal diameter, and
tiller and panicle production through the first two growing seasons, which was the
duration of subsequent monitoring (Richter & Stutz 2002). These results highlight the

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potential of AMF inoculation for restoration, the importance of focusing on factors other
than growth in determining success, and the need for long-term monitoring of
reintroduced plants. Interestingly this study also found that plants started in smaller
containers had greater survival, height, basal diameter, and tiller production (Richter &
Stutz 2002).
Hawaii is home to 41% of endangered plant species in the US (as of 2002) and
propagation in greenhouses and later reintroduction of seedlings to the field has been an
important and mostly unsuccessful element of conservation (Gemma et al. 2002). Four
endemic species, two of which are listed endangered, were grown with local AMF
inoculant, and growth of both roots and shoots was significantly enhanced in all species
tested compared to control plants in the greenhouse (Gemma et al. 2002). Growth was
especially enhanced in low-P soils, and P limitation is widespread in Hawaii (Gemma et
al. 2002). No outplanting or long-term survival data were included in this paper; however,
the authors suggest that the strong positive plant response to inoculation may indicate
greater potential for field success.
In an AMF inoculation trial of six native montane species from Rocky Mountain
National Park (CO), Rowe et al. (2007) found a significant though varying growth
response to inoculation. Three late-successional species showed a positive response to
AMF, while three early successional species showed a negative response (Rowe et al.
2007). Both native and commercial AMF inoculants were tested, and native-cultured
inoculant produced higher levels of root colonization and plant response (Rowe et al.
2007). Soil P levels had an effect on AMF responsiveness for only one plant studied
(Rowe et al. 2007).

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Three plant species, two of which are on the International Union for Conservation
of Nature (IUCN) Red List, meaning that they are of highest conservation priority, and
one of which is extinct in the wild in Poland, were tested for response to native AMF, a
mixture of laboratory AMF strains, and a combination of laboratory strains and
rhizobacteria in Europe (Zubek et al. 2009). Inoculation type did not have a significant
effect, but all three plants were shown to be dependent on AMF with two having extreme
19- to 22-fold and 11- to 14-fold gains in biomass over non-mycorrhizal controls (Zubek
et al. 2009). The authors conclude that AMF inoculation should be used in the
propagation of these species and that it is likely to aid in future success of outplanting and
reintroduction (Zubek et al. 2009).
In climate-change-related reintroduction research, Ferrazzano and Williamson
(2013) inoculated seeds of an endangered plant with AMF and planted the Abronia
macrocarpa seeds directly into plots in an area of Texas that was experiencing drought
(2013). Growth factors, including mean number of leaves and mean aerial diameter were
significantly greater in AMF-treated plants (Ferrazzano & Williamson 2013). This study
shows not only that AMF inoculation can aid A. macrocarpa, but that mycorrhizal
colonization of plants may become increasingly important under global warming.
The results of these five studies all show positive response of rare plants to AMF
inoculation for at least some of the species studied, and indicate that AMF inoculation is
a strategy that should be tried on more species, in different ecosystems, and with longterm survival monitoring. It is important to keep in mind, however, that a positive-results
bias exists in data reporting and publishing, and that these data are likely not indicative of
the total research that has occurred in studies of mycorrhizal inoculation and rare plant

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reintroduction (Heidorn 2008). This important research should continue, and both
positive and negative results should be shared publicly. Widespread sharing of all results
is becoming more possible as organizations increasingly host websites where
reintroduction success, failure, and methods data may be submitted anonymously and
accessed through databases (Guerrant 2012). In addition, land managers who decide to
test and use mycorrhizal inoculation should be aware of the complexities of both
mycorrhizal symbiosis and plant reintroduction and the need to collect and consider data
beyond initial growth rates, and that not all species will respond in the same way.

Inoculant source
It has been shown that different species of AMF affect plants differently, and that these
effects occur not only based on species, but genotype (Anderson & Roberts 1993;
Klironomos 2003). Whole soil containing inoculant collected from local sites is a
potential source of AMF that has been well-studied. Rowe et al. (2007) found
mycorrhizal colonization of 100% of plants treated with field soil and only 8.4% of those
that received a commercially available inoculant. Unfortunately local whole-soil
inoculants, while cost-effective, are not always available and can include unwanted
pathogens. In ecological restoration it can be especially problematic to remove soil from
intact sites, which could cause damage, especially if done regularly or in large amounts.
Degraded sites may contain few AMF or an altered species composition (Vogelsang &
Bever 2009). In some cases, soil from high-quality sites may be almost completely
unavailable, such as in the prairies of South Puget Sound, Washington, where the highest

27
 


 

 
quality remnant patches exist on a military base and digging is not allowed due to the
presence of unexploded ordinances.
Samples of AMF can be collected from the roots of native plants, isolated, and
added to sterilized media in “pot cultures” (IJdo et al. 2011). This method reduces the
need for field soil, and the possibility of unintentionally introducing pathogenic
organisms. It can, however, take years to produce sufficient inoculant for large-scale use,
and cultures have a tendency to become reduced in AMF species complexity over time.
The phenomenon is due to unintentional selection for AMF that thrive in a greenhouse
environment and are able to competitively colonize the plant species grown in cultures
(IJdo et al. 2011). Accidental contamination with non-native AMF species can also occur,
and is difficult to detect.
Commercial inoculants are a readily available and appealing source of AMF, and
contain species that are known to grow quickly and colonize a wide variety of plant
species. These “general” inoculants also have disadvantages. The quality is variable, and
difficult to test, and species that have the best initial colonization abilities in the
greenhouse may not be the most beneficial for long-term growth in the field (Schwartz et
al. 2006; Rowe 2007). Introducing non-native AMF species into the field is also a very
real, though not well-studied possibility (Schwartz et al. 2006). Plants inoculated with
AMF in the greenhouse tend to retain that AMF community identity, even after planting
in field soil (Mummey et al. 2009).

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Risk
Human-enhanced movement of plant, animal, and pathogen species around the
world through globalization has had devastating ecological, social, and economic
consequences. Many habitats have become severely degraded, leading to poor
functioning and reduced ecological services, and invasive species have been a major
force, along with habitat loss, in causing extinctions (Schwartz et al. 2006). While no
known problems from the introduction of non-native AMF have been documented,
Schwartz et al. (2006) note that this may be due to size. Human ability to notice and
record invasions tends to correlate with the size and therefore visibility of the invasive
organism, vertebrates tend to be quickly noticed, while very little is known about the
history of earth worm invasions, despite their effects on ecosystems in North America.
Exceptions to this rule of size do occur, fore example when larger species that we
consider important are visibly affected, such as with the diseases that caused the chestnut
blight or potato famine. AMF, however, are both invisible to the human eye and not welldocumented to begin with, leading to the very real possibility of negative effects of
invasion going unnoticed.
Introduction of non-native AMF could lead to both biological and chemical
changes to ecosystems with potentially global consequences. Unlike AMF,
ectomycorrhizal fungi have been known to become invasive and alter biogeochemical
cycles. In Ecuador, many highlands that once contained paramo grasslands have been
planted with non-native radiata pine (Pinus radiata) saplings that had been grown in soil
containing EMF inoculum from older pines (Chapela et al. 2001). Conversion of
grasslands to forest leads to an increase in carbon stored in aboveground biomass and has

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received attention as a mitigation strategy to combat global warming, under the
assumption that belowground C would not change. However, researchers found a loss of
up to 30% of soil C twelve years after plantation (Chapela et al. 2001). Through a variety
of techniques including stable isotope and EMF genetic analysis, Chapela et al. (2001)
found that EMF diversity was extremely low compared to native radiata pine forests, and
that the fungi were acting saprotrophically, effectively leading to “photosynthesis-derived
soil C mining.” Carbon from the soil C pool, established by grasslands and AMF, was
being turned into an abundance of fungal fruiting bodies, with productivity of Suillus
luteus alone up to three orders of magnitude greater than all EMF combined in
comparable native forests (Chapela et al. 2001). This study shows the drastic and
unexpected effects that non-native mycorrhizal fungi can have on an ecosystem and its
biogeochemical cycling.
AMF, regardless of nativity, may also benefit invasive plants. While AMF
densities have been shown to decline in association with invaders, and low densities
contribute to invasion, many common weeds have been shown to respond positively to
AMF (Vogelsang & Bever, 2009). AMF networks can also work against native plants, as
was found in an experiment where spotted knapweed (Centaurea maculosa), a
problematic invader, was shown to experience maximum growth benefit with a native
fescue (Festuca idahoensis) and AMF, in comparison to trials without AMF or with
another noxious weed, even though C. maculosa’s photosynthetic rate was 14% lower
than when it was grown alone (Carey et al. 2004).
While AMF inoculation does carry risks, it is important to remember that no
serious damaging effects have yet been documented, and that all management activities

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carry risk, as does a lack of management. AMF, and especially native cultured AMF are
likely much less risky than other practices such as pesticide use, and may well help to
save species from extinction.

Figure 4. Geographical extent of the Willamette Valley–Puget Trough–Georgia Basin Ecoregion.
Historically, much of this area was dominated by prairies and oak savannas maintained through
anthropogenic use of fire. Map from Hamman et al. 2011.

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PACIFIC NORTHWEST PRAIRIE-OAK ECOSYSTEMS
The prairie and oak savannas of the Willamette Valley–Puget Trough–Georgia Basin
(WPG) ecoregion are among the most endangered ecosystems in the United States
(Floberg 2004; Dunwiddie & Bakker 2011). This ecoregion is located in the Pacific
Northwest, and occurs between the Cascade Range and coastal mountains from southern
British Columbia to southern Oregon, USA (Fig. 5) (Hamman et al. 2011). While it is
difficult to quantify the original extent of these ecosystems, most experts agree that <10%
of these habitats remain, and that <5%–1% are considered “intact” and dominated by
native species (Dunwiddie & Bakker, 2011; Hamman et al., 2011). A variety of factors
have led to the disappearance and degradation of WPG prairies including non-indigenous
settlement, woody species encroachment, the introduction of non-native plants, and most
importantly, the extirpation of indigenous people and their land management practices
(Boyd 1999).
Overwhelming evidence, including historical narratives, charcoal and pollen
records, dendrochronology, plant physiology, ecosystem ecology, and indigenous
knowledge, indicate that WPG prairies and oak savannas were shaped and maintained by
the intentional use of fire (Cooper 1859; Boyd 1999; Hamman 2011; Sprenger et al.
2011; Walsh et al. 2010). Regular anthropogenic use of burning in the WPG is thought to
have ceased in the mid-nineteenth century when lands were heavily settled by EuroAmericans and converted to agriculture. Both historical records and charcoal evidence
suggest, however, that in many parts of the WPG region anthropogenic burning slowed or
ceased around 1700, due to indigineous mortality through disease, and that the landscapes
first viewed by settlers around 1850 were already heavily altered by 150 years of woody

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encroachment and Native American–population decline (Cooper, 1860; Walsh et al.,
2010). Restoration of the prairie and oak savannas of the WPG, and the species that
evolved and coexisted under indigenous use of high-frequency, low-severity fire, is the
preservation of not just a biological landscape, but a cultural one.
Several species that are endemic to WPG prairies have been listed or are
candidates for threatened and endangered status, according to the ESA including: the
hemiparisitic golden paintbrush (Castilleja levisecta), the Taylor's checkerspot butterfly
(Euphydryas editha taylori), Mardon skipper (Polites mardon), streaked horned lark
(Eremophila alpestris strigata), and Mazama pocket gopher (Thomomys mazama)
(Dunwiddie & Bakker, 2011). More than 100 species associated with WPG prairie and
oak savannas are considered “at risk” in British Columbia (Dunwiddie & Bakker 2011).
Forty-six plant species that are fire-adapted or fire-dependent and specific to the WPG
prairie and oak savannas are either critically imperiled, imperiled, or vulnerable
(Hamman 2011). Protection of biodiversity has made these prairie and oak savannas a
high priority in the region for both research and conservation efforts, which have been
underway and steadily maturing in both scope and technique over the last twenty years
(Dunwiddie & Bakker 2011).
Regional restoration efforts include the use of prescribed fire and species
reintroductions, and appropriately sourced seeds and seedlings are produced in large
quantities for reintroduction (Hamman 2011). Plant reintroduction efforts suffer from the
same problems with long-term establishment that plague restoration projects globally
(Hamman, personal communication). It appears that AMF have not been studied in the
WPG prairie oak savanna ecoregion, with two exceptions. The first is a master’s thesis by

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Sierra Smith that focused on the reconversion of former agricultural land to prairie (2007).
Achillea millefolium and Festuca scabrella were grown in soil with or without AMF in a
greenhouse, and negative growth effects occurred, though interestingly a trend toward
greater vigor in AMF seedlings was also found (Smith 2007). A second study in South
Puget Sound, WA, tested inoculation of several plants, but issues with seed germination
occurred and fertilizer was added to non-AMF control plants, making data comparison
problematic. Nevertheless, initial field results indicate that AMF-inoculated plants are
experiencing greater survival than their fertilized counterparts (Hamman, personal
communication).
WPG prairie plant species are excellent candidates for mycorrhizal inoculation
and long-term success, due to low native soil N and P, and other restoration efforts that
are occurring, including regular use of prescribed fire (Hamman, 2011). A large number
of groups including nonprofit conservation organizations, government agencies (federal,
state, and local), and Joint Base Lewis-McChord (the landholdings of which include the
largest remnants of intact prairie in the ecoregion), are heavily invested in WPG prairieoak restoration (CNLM, 2013). A portion of seed produced is dedicated to research
efforts, and the major financial and labor involvement of a wide-range of stakeholders
makes long-term monitoring of experimental efforts possible (Sarah Hamman, personal
communication, 2013). Research into the effects of AMF in the propagation of rare and
native WPG prairie plants will potentially address issues of both basic and applied
science at a variety of scales, and inform restoration and conservation efforts globally.

 


 

34
 


 

 
This Research Article has been formatted for submission to the journal Restoration
Ecology.

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Mycorrhizal and microbial inoculation affect the growth and survival of
native plants raised for restoration
Sasha R. Porter,1,2 Erin E. Martin,2 Sarah T. Hamman1
1

Center for Natural Lands Management, 120 Union Ave SE, Olympia, Washington 98501

2

The Evergreen State College, Graduate Program on the Environment, 2700 Evergreen

Parkway NW, Olympia, Washington 98505
Key words: arbuscular mycorrhizae, native plants, restoration, soil microbial community,
Willamette Valley–Puget Trough–Georgia Basin prairie, greenhouse propagation
Running title: Mycorrhizal and Microbial Inoculation
Implications for practice:
•

Use of arbuscular mycorrhizal fungi inoculants in the greenhouse can enhance
seedling growth and survival

•

An AMF inoculant cultured from the roots of local native plants worked as well
as a commercially available one

•

Care should be taken with whole-soil, rather than cultured inoculants, which may
contain pathogens and decrease seedling emergence, survival, and growth

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ABSTRACT
Production of native seedlings for field outplanting has become a common ecological
restoration technique worldwide. However, the establishment of greenhouse-raised plants
in the field is usually poor. Mycorrhizal fungi are symbionts that can provide survival
benefits to host plants. This relationship is ubiquitous in terrestrial ecosystems and
mycorrhizae are absent only under unusual circumstances, such as in a nursery
greenhouse.
In this study, nine Northwest short-grass prairie species were grown for six months in
sterilized medium with an arbuscular mycorrhizal fungi (AMF) inoculant cultured from
local native plants, a general AMF inoculant, or in control treatments. Three microbial
inoculants with AMF removed, created from a nearby site considered to be high-quality
remnant prairie, a restoration site, and unsterilized potting medium, were added within
each AMF treatment in a full factorial design. Seedling emergence, survival,
aboveground growth, and biomass data were collected.
AMF significantly enhanced the growth of five species and the survival of four, with no
detectable effect on the remainder. Further, there was no significant difference between
the two AMF inoculants. Field microbial wash tended to have a negative effect on
seedling emergence and growth, with the high-quality site treatment most repressive.
AMF and the introduced microrganisms interacted on Festuca roemeri, with AMF
mediating the negative effect of other fungi. Surprisingly, AMF positively affected the
growth of Castilleja levisecta, a hemiparasite, and altered the phenology of Dodecatheon
hendersonii, delaying dormancy. These results suggest that AMF can enhance the growth
and survivorship of many species, and that inoculation may lead to greater success in
ecosystem restoration efforts.

INTRODUCTION
Many scientists agree that we are in the midst of Earth’s sixth great extinction event,
characterized by the loss of important natural and cultural landscapes, and a steady
decline in biodiversity globally (Barnosky et al. 2011). Encouraging the re-establishment
of viable wild populations of rare and native plant species in degraded ecosystems
through cultivation and outplanting is a widespread restoration technique. Plant
reintroduction attempts to mitigate loss of biodiversity and to prevent extinction by
increasing native plant abundance and diversity, with a resultant preservation of species

37
 


 

 
across trophic levels. The actual success rate of plant reintroductions, defined by
Godefroid et al. (2011) as the ability of a population to persist and reproduce, however, is
quite low; long-term monitoring of the outcome of reintroduction efforts is infrequent,
and the authors found that published literature reflects a strong bias toward successful
results (78%), as compared to those reported in survey data (33%). Where longer-term
results were available, a startling decline in success occurred over time with an average
of only 6% of reintroduced plants alive and flowering after 4 years (Godefroid et al.
2011).
The failure of greenhouse-raised transplants is due in part to horticultural
techniques that provide abundant nutrients but poorly emulate native environments
(Haskins & Pence 2012). Field soils contain complex microbial communities capable of
altering plant growth and survival while growing media often lacks biota or contains
communities that are compositionally distinct from those in outplanting sites. Lack of
mycorrhizal fungi, symbiotic partners with which plants evolved, may be especially
problematic for greenhouse-raised perennial seedlings (Haskins & Pence 2012).
Arbuscular mycorrhizal fungi (AMF) consist of mycelium formed by masses of
branching, threadlike hyphae that enter roots and receive carbon from the host plant in
exchange for extended access to resources; it is through AMF, not roots, that the majority
of nutrients are taken up by plants (Smith & Read 2008). AMF have been shown to affect
plant growth, survival, diversity, hormone levels, morphology, tolerance to toxins,
herbivory, disease, and drought (Hartnett and Wilson 2002, van der Heijden 2004, Smith
& Smith 2012). These effects are influenced by factors such as plant species, age,
community composition, AMF species and diversity, resource levels, edaphic biota, and

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feedbacks between these and other complex elements (Klironomos 2003, Hoeksema et al.
2010, Smith & Smith 2012). Some plants cannot survive without colonization, while
others, especially annuals and early successional species, may show a negative growth
response to AMF because plant carbon is traded for non-limiting nutrients (Smith &
Smith 2012). Degraded environments often contain less AMF, so inoculation in the
greenhouse could give transplants not only an equal playing field, but an advantage over
neighbors (Vogelsang & Bever, 2009).
Previous greenhouse studies of AMF effects on the propagation of native, rare,
and endangered plants have shown positive effects on four species in Hawaii (Gemma et
al. 2002), two forbs in Florida (Fisher & Jayachandran 2002), a grass in Arizona (Richter
& Stutz 2002), three late-successional species in subalpine Colorado (Rowe et al. 2007),
and three montane species in Poland (Zubek et al. 2009). Seedlings were outplanted in
only one of these studies; for the study period examined (two growing seasons) Richter
and Stutz (2002) found that AMF inoculation led to increased survival, basal area, and
tiller and panicle production. While the current literature indicates positive effects,
implementation in conservation lags behind agriculture and horticulture, where use is
more common (Haskins & Pence 2012).
For ecological and economic reasons conservationists may be especially
concerned with the source of mycorrhizal inoculants, which can be produced from native
ecosystem-specific AMF strains, sourced from generalized commercially available AMF
communities, or introduced by adding small amounts of whole-soil to growing medium
(Schwartz et al. 2006). While AMF have not been known to become invasive, Schwartz
et al. (2006) recommend using inoculants cultured from local sources to avoid

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introducing nonnative strains to field environments. Further, individual plant species have
growth responses that vary from highly positive to negative based on AMF species
identity and source, with local genotypes producing the strongest reactions (Klironomos
2003). In addition, the presence of other microorganisms, such as bacteria, tends to
increase plant response to AMF through complex interactions including increased
nutrient cycling and the presence of mycorrhization helper bacterium (Hoeksema et al.
2010, Kurth et al. 2013). We hypothesized that native cultured AMF combined with a
microbial community from a high-quality prairie site would lead to greater plant growth
and survival than a general AMF inoculant, or microbes from other sources.
Prairie and oak savannas of the Willamette Valley–Puget Trough–Georgia Basin
(WPG) ecoregion in the Pacific Northwest are among the most endangered ecosystems in
the United States with <10% remaining, and only between 1-5% dominated by native
species (Dunwiddie and Bakker 2011, Hamman et al. 2011). A variety of factors have led
to the disappearance and degradation of WPG prairies including increased settlement and
agriculture, invasion by trees and non-native plants, and the extirpation of indigenous
people and their land management practices, which included the regular application of
fire. US Endangered Species Act (ESA)–listed organisms in the area include the
hemiparisitic golden paintbrush (Castilleja levisecta), Taylor's checkerspot butterfly
(Euphydryas editha taylori), streaked horned lark (Eremophila alpestris strigata), and
Mazama pocket gopher (Thomomys mazama). Further, 46 plant species that are fireadapted or -dependent and specific to the WPG are either critically imperiled, imperiled,
or vulnerable (Hamman, 2011).

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This study addressed the use of AMF inoculation for the production of restoration
seedlings by testing nine native WPG species with two AMF inoculants and three
microbial cultures. A secondary goal was to produce sufficient material for outplanting
and long-term study. The following questions were addressed 1) How does AMF
inoculation affect the growth of greenhouse-raised seedlings? 2) Does AMF affect shortterm (6 months) survival? 3) Is an AMF inoculant cultured from native soils superior to a
commercially available one? 4) How will AMF inoculants interact with different soil
microbial communities likely to be present in outplanting sites?

MATERIALS AND METHODS
Seed stratification and sowing
Seeds of Aquilegia formosa Fisch. ex DC., Balsamorhiza deltoidea Nutt., Castilleja
levisecta Greenm., Dodecatheon hendersonii A. Gray, Dodecatheon pulchellum (Raf.)
Merr., Festuca roemeri (Pavlick) E. B. Alexeev, Gaillardia aristata Pursh, Micranthes
integrifolia (Hook.), Ranunculus occidentalis Nutt., and Silene douglasii Hook. were
imbibed with tap water in filter paper inside a pipet washer (Thermo Scientific, Waltham,
MA) for 0–24 hours (Supp. Fig. 1, Table 1, Supp. Table 1). Seeds were then placed in
sterilized petri dishes on moist filter paper and underwent stratification in a dark
environmental chamber at 3° C for 0–90 days (Supp. Table 1). Imbibition and
stratification times, where available, were based on previous research (Krock et al. in
review).

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Species
 

Common
 Name
 

Family
 

Status
 

Other
 

Aquilegia
 formosa
 

Western
 columbine
 

Ranunculaceae
 
Asteraceae
 

G5,
 S2
 


 
E.
 taylori
 nectar
 

Balsamorhiza
 deltoidea
 

Deltoid
 balsamroot
 

Castilleja
 levisecta
 

Golden
 Paintbrush
 

Orobanchaceae
 

G1,
 S1
 

E.
 taylori
 ovipositon
 

Dodecatheon
 hendersonii
 

Mosquito
 bills
 

Primulaceae
 


 

Dodecatheon
 pulchellum
 

Darkthroat
 shootingstar
 

Primulaceae
 

Festuca
 roemeri
 

Roemer’s
 fescue
 

Poaceae
 


 


 
P.
 mardon
 ovipositon
 

Gaillardia
 aristata
 

Blanket
 flower
 

Asteraceae
 


 

Micranthes
 integrifolia
 

Whole-­‐leaf
 saxifrage
 

Saxifragaceae
 


 
E.
 taylori
 nectar
 

Ranunculus
 occidentalis
 

Western
 buttercup
 

Ranunculaceae
 

E.
 taylori
 nectar
 

Silene
 douglasii
 

Douglas
 campion
 

Caryophyllaceae
 


 

Table 1. Ten native northwest prairie species were selected for this study. Status indicates
endangered ranking globally (G) and by state (S) with lower numbers signifying most critically
imperiled. “Other” shows known ecological benefit to the endangered Taylor’s Checkerspot butterfly
(Euphydryas editha taylori) or Mardon Skipper (Polites mardon).

Growing medium and inoculants
Medium for all treatments was Sunshine mix #2 (Sun Gro, Agawam, MA), a mix of
coarse Canadian sphagnum peat moss, perlite, and dolomite. The soil was loosened and
then sterilized in an autoclave at 121° C on gravity cycle for 50 min in 15 L batches. To
simulate native prairie soil nutrient ratios, Apex 16-5-9 NPK Plus slow release fertilizer
(J.R. Simplot, Boise, ID) was added at a ratio of 3 g per L of medium, the manufacturer’s
lowest recommended level.
Seeds were grown in a greenhouse with one of four arbuscular mycorrhizal fungi
(AMF) inoculation treatments: 1) Native (NA) inoculant, 2) General (GE) inoculant, 3)
Control (CO—no inoculant), 4) Fungicide treatment (FU—additional fungal control).
The fungicide Thiophanate-methyl (dimethyl 4,4’-o-phenylenebis[3-thioallophanate])
(50% a.i., Cleary Chemical, Dayton, NJ), which has been shown to suppress mycorrhizal
colonization, was added to the soil of the FU treatment at the rate of 50 mg (active

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ingredient) per kg medium (dry mass) (Wilson & Williamson 2008). Contamination of
controls in AMF experiments is a potential problem that can be greatly reduced by adding
fungicide, however the addition may cause unwanted side effects through the elimination
of other fungi present in AMF treatments, especially when a microbial wash has been
applied (Wilson and Williamson 2008). We used both a non-AMF control and a
fungicide control to allow comparison between the two.
Both NA and GE mycorrhizal inoculants were cultivated by Plant Health, LLC
(Corvallis, OR). The NA inoculant was cultivated from the roots of eight native plants on
intact prairie sites at Joint Base Lewis McChord (JBLM). Fifty grams of inoculant were
added per L of growing medium in the NA and GE treatments (personal communication,
B. Linderman).
Autoclaving eliminates not only mycorrhizal fungi, but all edaphic
microorganisms. To simulate a variety of more-realistic soil conditions we created three
soil microbial washes, with mycorrhizal spores removed, from 1) High-quality prairie (HI,
>24 native plant species, JBLM), 2) Restoration prairie (RE, <10 native species, The
Nature Conservancy), and 3) A control of potting soil (PO). Each microbial wash was
applied to 1/3 of each mycorrhizal treatment in a full factorial design to elucidate
interaction effects between AMF and soil microbial communities (Supp. Fig. 2).
Following the methods of Koide and Li (1989), 3 mL deionized water per gram of field
or unsterilized potting medium was combined to form a slurry, which was then filtered
through a 38µm sieve to remove AMF spores. The resulting filtrate was applied at 30 mL
microbial wash per L sterilized potting medium (Koide & Li 1989).

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Experimental design
Each species was sown into 384 plugs (Ray Leach Cone-tainer, Tangent, OR), which had
previously been sterilized with a 10% bleach solution, with two to three seeds per 107
mL plug. Plugs had been filled with growing medium that received one of four AMF
treatments (NA, GE, FU, CO) and one of three microbial treatments (HI, RE, PO), for a
total of 12 treatment combinations (plugs n = 3456; mycorrhizal treatment n = 864;
microbial treatment n = 1,152; mycorrhizal x microbial n = 288; mycorrhizal x microbial
x species n = 32). Flats were placed together in threes among 12 groups that each
contained all nine species with the same AMF treatment and microbial wash; 20 cm
separated groups to reduce the possibility of cross-contamination (Supp. Fig. 2). At the
time of planting seeds were covered with sterilized potting soil equal to two times the
thickness of the seed to reduce the possibility of cross-contamination between trays
(Grman 2012). A thin layer of sterilized gravel (Gran-I-Grit starter, Mount Airy, NC) less
than 3 mm deep was also added to the top of each plug to stabilize seeds and soil, and
reduce bryophyte growth. Propagules were watered daily with an overhead sprinkler
system and trays were rotated in the unheated greenhouse weekly.
All species were sown in February 2014, except for F. roemeri, which was
planted in April 2014 to replace A. formosa, which had experienced less than 3%
germination. Plugs of A. formosa containing seedlings were removed and remaining
ungerminated plugs sown with 5–7 seeds of F. roemeri, selected due to its quick and
consistent germination, its role as an ecosystem keystone species, and to include an
additional family in the study. Any difference that occurred due to ungerminated A.
formosa seeds was consistent throughout F. roemeri treatments.

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Data collection and analyses
Germination data were collected weekly until, at nine weeks (when emergence had
slowed and die-off begun) all plugs without seedlings were removed and the remainder
thinned to the largest seedling per plug. This allowed the subsequent monitoring of
survival of the one remaining individual rather than germination. Survival was tracked
weekly for ten weeks and bi-monthly thereafter. Height and/or width data were collected
at 16 weeks based on plant morphology, with procumbent species measured for width,
upright species for height, and R. occidentalis for both.
At six months 60 plants from two species that had sufficient replicates (D.
hendersonii, M. integrifolia) for long-term outplanting were harvested for biomass
analysis. The first four plugs with visible aboveground growth were selected from each
subgroup (mycorrhizal x microbial) for M. integrifolia. For D. hendersonii, which was
dormant in all treatments, the first four plugs were harvested. Shoots were cut from roots
at growing medium level, shoots and roots were washed with tap water, and dried at 60°
C for 48 hours. Roots and shoots were weighed separately using an analytical balance
(Mettler-Toledo, Switzerland).
All statistical analyses were completed using JMP Pro 11.0 software (SAS Inc.,
Cary, NC). Data were first analyzed for homogeneity of variance using Levene’s test, and
all data met requirements. Two-way ANOVA was used to analyze interaction effects of
AMF and microbial wash. For height, width, and biomass, absent plants were eliminated
from data. Chi-square was used for seedling emergence and survival analysis. Date of
peak die-off was used for species that had experienced significant or complete die-back,

45
 


 

 
most-recent data for all others. Tukey’s HSD with an alpha level of 0.05 was used to
compare means.
The two mycorrhizal treatments were combined after preliminary analysis showed
no statistically significant difference between NA and GE AMF inoculant for any species.
As there was also no difference between the CO and FU treatments, except for in the case
of F. roemeri, these treatments were also binned for all other species, allowing greater
statistical power and simplicity of data presentation. For F. roemeri CO and FU data were
not combined, and ANOVA of each AMF x microbial group was executed.

RESULTS
Growth
At 16 weeks, inoculation with arbuscular mycorrhizal fungi had a significant positive
effect on the growth of the majority of plants tested, including Castilleja levisecta
(F[1,77]=5.75, p<0.02), Dodecatheon pulchellum (F[1,292]=19.23, p <0.0001), Micranthes
integrifolia (F[1,351]=159.35, p <0.0001), Ranunculus occidentalis (height F[1,113]=4.05, p
<0.05, width F[1,113]=5.15, p <0.03), and Silene douglasii (F[1,77]=5.93, p <0.02) (Fig. 1).
AMF inoculation did not significantly affect the growth of Balsamorhiza deltoidea
(F[1,43]=0.06, p=0.81), Dodecatheon hendersonii (F[1,67]=2.5, p=0.19), Festuca roemeri
(see below), or Gaillardia aristata (F[1,39]=3.79, p=0.06) (Fig. 1).
 
There was no significant difference between the NA AMF inoculant and GE
inoculant on growth at 16 weeks for any species (data not shown). There was also no
difference between the CO and FU control treatments, except for on F. roemeri, where
the FU treatment significantly enhanced growth (F[2,348]=13.45, p<0.0001) (Fig. 2).

46
 

!
!

!

!

Figure 1. Growth response of seedlings to inoculation with arbuscular mycorrhizal fungi at 16 weeks.
Dark gray bars show AMF treatment, white control, and light gray fungicide. Height, width, or both
were measured based on species morphology. For F. roemeri, fungicide and control treatments were
different and both are shown, lower case letters denote significant differences (Tukey’s HSD) at an
alpha level of 0.05. Bars show ± one standard error from the mean.
*indicates statistically significant effect

47!


 

 
!

CONTROL p = 0.01*

AMF p = 0.79

FUNGICIDE p = 0.09

10"
9"
8"

height&(cm)&

7"
6"
5"
4"
3"
2"
1"
0"

CO"HI"

CO"RE"

CO"PO"

MICROBIAL SOURCE:

AMF"HI"

AMF"RE"

High-quality

AMF"PO"

Restoration

FU"HI"

FU"RE"

FU"PO"

Potting soil (control)

!


 
Figure 2. ANOVA showed that arbuscular mycorrhizal fungi (AMF) mediated the negative effects of
microbial wash on Festuca roemeri. AMF treatments are grouped, CO=control, AMF=inoculant,
FU=fungicide (Thiophanate-methyl), within, microbial treatments are black HI=high-quality prairie,
gray RE=restoration site, white dotted PO=potting soil. Bars show ± one standard error from the
mean.


 
Both species harvested for biomass measurements had significantly greater root
biomass with AMF inoculation (D. hendersonii F[1,44]=10.21, p<0.003, M. integrifolia
F[1,45]=29.67, p<0.0001) (Fig. 3). Shoots were only available for M. integrifolia, as D.
hendersonii had entered dormancy, and were also positively affected by AMF inoculation
(F[1,45]= 20.07, p<0.0001), as was total biomass with D. hendersonii showing a greater
than 70% increase with AMF (F[1,44]=10.21, p<0.003) and M. integrifolia 189%
(F[1,46]=32.32, p<0.0001). Shoot-root ratio was not significant (F[1,46]=0.35, p=0.56) (Fig.
4). The mass of individual D. hendersonii roots showed a significant correlation to week
entering dormancy with longer-living shoots producing larger roots (r2=0.35,
F[1,40]=21.27, p<0.0001) (Fig. 5).
 

48
 


 

 

Figure 3. Roots of Micranthes integrifolia harvested at six months from plants A) in a
nonmycorrhizal control and B) inoculated with arbuscular mycorrhizal fungi.


 

0.6
 

D.
 
hendersonii*
 
0.01
 

biomass
 (grams)
 

biomass
 (grams)
 

M.
 
integrifolia*
 

0.4
 
0.2
 
0
 

0.005
 

0
 

p
 <
 0.0001
 

p
 <
 0.003
 


 

Figure 4. Belowground biomass of Micranthes integrifolia and Dodecatheon henderonii were
significantly increased by inoculation with arbuscular mycorrhizal fungi (gray, control white). Bars
show ± one standard error from the mean.


 


 

49
 


 

 


 
Figure 5. Correlation of root biomass and week entering dormancy for Dodecatheon hendersonii.
Plants were harvested six months after planting. The line shows the relationship between week
entering dormancy and root biomass for all treatments.

Two species showed a significant growth response to microbial inoculant with the
restoration treatment (RE) suppressing B. deltoidea (F[2,42]=4.51, p <0.02), and the highquality treatment (HI) suppressing R. occidentalis (height F[2,112]=6.58, p=0.002, width
F[2,112]=5.15, p=0.007) (Fig. 6). Two-way ANOVA showed significant interaction
between AMF inoculant and microbial inoculant on growth for F. roemeri
(AMF*microbial F=4.83, p<0.009) (Table 2).

Source
 

df
  Sum
 of
 squares
 

F
 

p
 

Model
 
Error
 

5
 
345
 

39.1
  3.45
  <0.005
 
782.95
 

 

 

Corrected
 total
 

350
 

822.06
 
 


 

Mycorrhizal
 

1
 

6.02
  2.65
 

0.1
 

Microbial
 

2
 

11.17
  2.46
 

0.09
 

Mycorrhizal*Microbial
 

2
 

21.94
  4.83
  <0.009
 

Table 2 Results of two-way ANOVA were significant for interaction between arbuscular mycorrhizal
fungi and microbial source for Festuca roemeri.

50
 


 

 

2!
0!

b!

4!
2!
0!

M.#integrifolia#

width&(cm)&

height&(cm)&

7!

1.5!
1!
0.5!

1!
0.5!

8!
6!
4!
2!
0!

p"="0.35"

p"="0.44"

D.#pulchellum#
2!

width&(cm)&

height&(cm)&

S.#douglasii#
10!

0!
p"="0.19"

6!
4!
2!
0!

1.9!
1.8!
1.7!
1.6!
1.5!

p"="0.82"

!

0.5!

p"="0.06"

1.5!

2!

p"="0.1"

1!

0!

C.#levisecta#

0!

G.#aristata#

2!

1.5!

p"="0.02"

2.5!

7.5!

4!

p"="0.007"

8.5!
8!

b!

0!

p"="0.002"

F.#roemeri#

a!

width&(cm)&

b!

a! a!

6!

D.#hendersonii#

height&(cm)&

4!

a!

6! a,b!

height&(cm)&

a,b!

B.#deltoidea*&

height&(cm)&

height&(cm)&

6!

R.#occidentalis#
width*&
width&(cm)&

R.#occidentalis#
height*&

p"="0.93"

!

Figure 6. Growth response of seedlings at 16 weeks to inoculation with three microbial washes (high-quality
prairie source dark gray, restoration gray, white control). Height, width, or both were measured based on
species morphology. Lower case letters denote significant differences (Tukey’s HSD) at an alpha level of 0.05,
bars show ± one standard error from the mean.
*indicates statistically significant result

51
 


 

 
Within F. roemeri mycorrhizal treatments there was a significant difference for
microbial wash within the CO treatment (F[2,86]=4.43, p=0.01) but not the AMF
(F[2,175]=0.24, p=0.79) or FU treatment (F[2,81]=2.4, p=0.09) (Fig. 2).

Emergence and Survival
Seedling emergence was unaffected by AMF for all but B. deltoidea, which showed a 7–
8% decrease in emergence with inoculation (X2[1,384]=4.83, p<0.03). Field microbial
inoculant tended to have a negative effect on emergence (Table 3a). This was observed
for the HI inoculant in B. deltoidea (X2[2,384]=14.65, p=0.0007), C. levisecta
(X2[2,384]=14.01, p=0.0009), G. aristata (X2[2,384]=13.63, p=0.001), and R. occidentalis
(X2[2,384]=22.4, p<0.0001), though there was a positive effect on D. pulchellum
(X2[2,384]=16, p=0.0003). The RE inoculant negatively affected B. deltoidea
(X2[2,384]=14.65, p=0.0007), and R. occidentalis (X2[2,384]=22.4, p<0.0001), but had a
positive effect on D. pulchellum (X2[2,384]=16, p=0.0003) and G. aristata (X2[2,384]=13.63,
p=0.001).
Survival of plants overall was significantly positively affected by AMF
inoculation (X2[1,1529]=18.5, p<0.0001). AMF did not have a negative effect on the
survival of any species and positively affected C. levisecta (at week 24 X2[1,85]=16.48,
p<0.0001), D. hendersonii (at week 15 X2[1,178]=0.003), M. integrifolia (at week 28
X2[1,314]=31.35, p<0.0001), and R. occidentalis (at week 28 X2[1,116]=3.78, p<0.05).
Microbial inoculation only affected the survival of F. roemeri, with the HI inoculant
having a negative effect (week X2[2,356]=6.17, p<0.05).

52
 


 

 


 
a)
 Seedling
 
 
Emergence
 
 
AMF
 
p
 

Balsamorrhiza
 
deltoidea
 

Castilleja
 
levisecta
 

Dodecatheon
 
hendersonii
 

Dodecatheon
 
pulchellum
 

Festuca
 
roemeri
 

Gaillardia
 
aristata
 

Micranthes
 
integrifolia
 

Ranuculus
 
occidentalis
 

Silene
 
douglasii
 


 


 


 


 


 


 


 


 

-­‐
 

0
 

0
 

0
 

0
 

0
 

0
 

0
 

0
 

<0.03
 

0.7
 

>0.05
 

0.7
 

*
 

0.46
 


 

0.08
 

0.13
 

HI
 micro
 

-­‐
 

-­‐
 

0
 

+
 

0
 

-­‐
 

0
 

-­‐
 

0
 

RE
 micro
 

-­‐
 

0
 

0
 

+
 

0
 

+
 

0
 

-­‐
 

0
 

<0.001
 

<0.001
 

0.2
 

<0.001
 

*
 

0.001
 

0.56
 

<0.0001
 

0.1
 

p
 
b)
 Survival
 


 


 


 


 


 


 


 


 


 

week
 

19
 

24
 

15
 

28
 

28
 

28
 

28
 

28
 

28
 

0
 

+
 

+
 

0
 

0
 

0
 

+
 

+
 

0
 

1
 

<0.0001
 

<0.003
 

0.06
 

0.63
 

0.76
 

<0.0001
 

0.05
 

0.83
 

HI
 micro
 

0
 

0
 

0
 

0
 

-­‐
 

0
 

0
 

0
 

0
 

RE
 micro
 

0
 

0
 

0
 

0
 

0
 

0
 

0
 

0
 

0
 

0.23
 

0.14
 

0.81
 

0.81
 

<0.05
 

0.17
 

0.43
 

0.6
 

0.27
 

AMF
 
p
 

p
 


 

Table 3. Chi-square analysis of the effects of arbuscular mycorrhizal fungi (AMF) inoculation and microbial wash on a) seedling emergence and b) survival. Effect
of inoculant is indicated as + (positive) or – (negative) where results are statistically significant, or 0 where there was no significant effect. HI = microbial wash
derived from high-quality intact prairie RE = microbial wash from a lower-quality prairie site undergoing restoration. Date of peak die-off was used in survival
analysis for species that had experienced significant or complete mortality, most-recent data for all others.
* F. roemeri experienced 100% emergence in all treatments

53
 

DISCUSSION
AMF Inoculation
Overall, inoculation with arbuscular mycorrhizal fungi (AMF) had a positive effect on
the growth and survival of the prairie species studied. As none of the species tested had a
negative growth or survival response to inoculation, and AMF will be present in
outplanting sites, inoculating all prairie species raised for outplanting without conducting
further species-specific trials, which can be expensive and time-consuming, would be
justifiable for restoration purposes.
We had hypothesized that plants grown with a native (NA) AMF culture would be
more likely to experience positive responses to inoculation, but found that a general (GE)
commercially available culture was equally beneficial, producing the same results for all
species. Acquiring, growing, and maintaining native field AMF cultures can be
problematic because: 1) field inoculum may be unavailable, 2) it is unknown whether
AMF field communities are representative of historical counterparts, and 3) for a variety
of reasons, AMF cultures tend to change over time (IJdo et al. 2011). The latter 2 cases
are likely true for this study, as individual AMF identity within the NA and GE cultures
has not yet been established, or the advantages of the NA culture may not be quantifiable
over the short term. Because it has been well-established that different plant and AMF
species, and even genotypes interact differently, and because introducing new nonnative
strains of AMF may be of concern (Schwartz et al. 2006), we believe that more research
on the development, maintenance, and effect of native AMF cultures for restoration is
needed, and that when possible practitioners should use native cultures (Anderson and
Roberts 1993, Klironomos 2003).


 

 
The significant positive response of over half of the species tested to AMF
inoculum after only six months indicates that AMF have likely been an important part of
the WPG ecosystem, allowing increased access to nutrients in gravelly glacial outwash
soils, and influencing the evolutionary strategy of plants that evolved there since the last
ice age. Vogelsang and Bever (2009) found that in California grasslands non-native
plants, which tend to depend less on AMF, cause mycorrhizal densities in soil to decrease,
creating less hospitable conditions for native forbs and a feedback loop that increases
invasion. A recent study on WPG prairies found that R. occidentalis roots were three
times more colonized by AMF in a high-quality remnant than in a more heavily degraded
restoration area (Block & Hamman, unpublished data). On sites that have already been
heavily affected by altered land use histories and non-native plant invasion (especially
shrubs and trees with ectomycorrhizal associates), AMF communities are likely already
in a degraded state and the reintroduction of mycorrhizal fungi may be beneficial
regardless of origin. However, managers should take more caution in outplanting
seedlings with potentially non-native AMF into intact ecosystems, though this suggestion
is based in caution rather than empirical evidence (Schwartz et al. 2006). Mycorrhizal
response of natives, which is often high, versus invasives, should be taken into
consideration before inoculating seedlings (Wilson & Hartnett 1998).
The only negative response to AMF found was a 7–8% decrease in Balsamorrhiza
deltoidea seedling emergence. Germination was problematic for several of the species in
this study, and may have been due to uncharacteristically cold conditions that reached
less than –9° C the first week after planting (NOAA 2014), the use of seeds that were ∼3
years old, or that some species had been selected for general study because of the desire

55
 


 

 
to improve the historically low germination rates. AMF are not generally thought to
affect seedling emergence, which was the case for eight of the nine species studied,
though Richter and Stutz found that inoculation positively affected Sporobolus wrightii
(2002). Poor germination rates combined with the goal of producing sufficient material
for a long-term outplanting study hindered the harvesting of most species for biomass
data, and we recommend that practitioners interested in studying AMF response overseed
at a rate of more than two to three seeds per plug for species that are known to have low
emergence rates. Heavily overseeding, however, should be avoided, as it may mask
results by producing germinants in every plug, as was seen with F. roemeri—it was the
variability in germination in this study that allowed us to analyze the negative effect of
field soil microbial washes on emergence (Table 3a).

Soil Microbial Wash
It is common practice in AMF experiments to reintegrate soil microbial communities
(sans AMF) back into sterilized media, and it has generally been found that plants
respond more positively to AMF with greater rhizosphere community complexity
(Hoeksema et al. 2010). We therefore expected the microbial wash that we derived from
a high-quality intact prairie site to have a positive interaction with AMF and on the
growth and survival of plants, but found instead that the HI and RE treatments tended to
repress plant growth, though reactions were species-specific and mostly not significant
(Fig. 2). The effect of field inoculant on seedling emergence, however, showed a negative
trend (Table 3a). A likely explanation for these effects is that field soils contained
pathogens that had differing species-based effects, though other causes cannot be

56
 


 

 
excluded. While the three washes likely contained different concentrations of
microorganisms, these concentrations were representative of sites and consistent within
treatments.
If pathogens were the main factor, plant survival was unaffected except for in F.
roemeri, and the repressive trend on growth could actually lead to increased field survival
as pathogen inoculated plants may have developed defenses under favorable greenhouse
conditions while PO plants will first be exposed in the competitive field environment
(Table 3b). The negative effect of field microbial inoculant on seedling emergence for
four species is concerning, suggesting that even the most intact remnant prairie sites may
be inhospitable places for native regeneration. Our data indicate that practitioners should
not apply microbial washes in the greenhouse until several weeks after seedlings have
emerged, and that the use of whole soil for AMF inoculation may be problematic in some
ecosystems even though it has been shown to have greater efficacy than cultured ones in
others (Rowe et al. 2007).

Species-Specific Responses
Dodecatheon hendersonii is an ephemeral species, one of the first flowers to appear on
WPG prairies in the spring and also one of the first to die back. During the first growing
season the species produces only cotyledons aboveground, concentrating energy into its
fleshy lateral root. At four months a significantly higher number of non-mycorrhizal
plants had entered dormancy than those treated with AMF (Table 3b). Unfortunately we
had not measured aboveground growth before the majority of this die-off occurred, which
likely skewed growth data through the measurement of only the few most vigorous plants

57
 


 

 
in the control treatments; if measurements had been taken earlier it is likely that growth
would have been significantly positively influenced by AMF, as well (Fig. 1). We
hypothesize that the extension of plants’ photosynthetically active phase would lead to
larger, more-developed belowground parts, and increased long-term survival. Data on
individual plant presence were collected weekly, allowing an analysis of the relationship
between week entering dormancy and biomass of belowground growth in plugs that were
harvested two months later. A strong relationship was found between week entering
dormancy and belowground biomass, with increasing aboveground survival time
correlated with larger root biomass (Fig. 6). In the field, plants that experience a longer
growth period may be more likely to experience pollination and establish viable
populations. We are currently unaware of other studies on the relationship between AMF
and plant phenology.
Another unexpected finding was the positive effect of AMF on the growth and
survival of Castilleja levisecta, a facultative hemiparasite. Parasitic plants produce
haustoria, physical structures that penetrate neighboring root systems, and have long been
considered some of the few non-mycorrhizal plants, however recent studies have shown
that AMF can colonize the roots of these species (Atsatt, Peter R. 1973, de Vega et al.
2010, Li et al. 2012). Parasitic plants have been observed to benefit from the AMF status
of their hosts, and may utilize the pathways that mediate AMF recognition and
colonization in host plants, causing hosts with higher AMF dependency to be more
receptive to parasites (Davies & Graves 1998, Fernández-Aparicio et al. 2010). While the
increased growth and survival of C. levisecta with AMF and no host is both surprising
and promising, it is unclear whether these benefits will continue under field conditions.

58
 


 

 
Li et al. (2012) found that a hemiparasite grown with AMF experienced reduced
haustoria formation and repressed growth after outplanting. This study provides an
excellent opportunity to further study the relationships between AMF, parasitic plants,
and hosts.
F. roemeri showed an especially interesting response to our treatments that
supports the theory that field microbial washes carried pathogens. This grass was the only
species for which the FU treatment, designed as a secondary control for AMF
contamination, had a significantly different effect than the CO treatment (Fig. 1).
Fungicide enhanced the growth of F. roemeri, likely by killing rusts (puccinia spp.),
which the plant is known to be susceptible to (Darris 2005), in the microbial wash. There
was no significant effect of microbial wash within the FU treatment, while the CO
treatment was negatively affected by the HI edaphic community (Fig. 2). Interestingly,
AMF inoculated plants were also not affected by microbial wash, though seedlings were
smaller than those that received fungicide, likely due to carbon allocation to AMF (Smith
& Read 2008). This indicates that AMF mediated the negative effects of the microbial
pathogen, though with a greater overall growth, but not survival, cost than fungicide.

Conclusion
These results indicate that use of AMF in the propagation of native species for
outplanting is likely to produce a greater number of seedlings overall, and further that
these seedlings will generally have larger above- and below-ground structures.
Practitioners should keep in mind that even where AMF inoculation does not positively
affect growth, it can act as a sort of “insurance policy”; it is advantageous for plants to

59
 


 

 
give small regular amounts of carbon, which can reduce short-term resources, for a
possible future payout if conditions become adverse, such as can occur with pathogens or
drought (Ferrazzano & Williamson 2013).
Upon outplanting, greenhouse propagules will be exposed to conditions that
include competition with neighboring native and non-native plants, exposure to
pathogens and other microorganisms, soil with different nutritional and physical
characteristics, herbivory from new species, and more-intense temperature and water
variability. The fact that plants are often linked by roots into complex mycorrhizal
networks with nutrients and signals transferred between plants may provide additional
benefits to plants in the field (van der Heijden & Horton, 2009). These interacting factors,
and changing environmental conditions, make it difficult to propagate greenhouse
seedlings that will be well-adapted to the field, and it is therefore important for
practitioners to rethink traditional horticultural techniques when producing plants for
restoration.
After six months of greenhouse study, it was found that AMF inoculation
successfully produced a greater number of plants with increased robustness, can extend
phenology, cause growth and survival benefits even in a hemiparasite, and mediate the
effects of soil pathogens. While this study focused on plants from one ecosystem, we
hope that results will encourage practitioners worldwide to try growing seedlings with
AMF, especially in areas characterized by low-nutrient soil or non-native invasion.
Seedlings produced in this experiment were outplanted in fall of 2014, and we anticipate
that field results collected over the long term will provide further answers to the questions
raised in this study.

60
 


 

 
ACKNOWLEDGEMENTS
Financial support for this research was provided by a Washington State Department of
Agriculture nursery research grant, and The Evergreen State College Office of
Sustainability. We also wish to thank Sierra Smith, Carl Elliott, Tel Vaughn, CNLM
interns, and the Evergreen Science Support Center.

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plant species reintroductions? Biological Conservation 144:672–682.
Hamman ST, Dunwiddie PW, Nuckols JL, McKinley M (2011) Fire as a restoration tool
in Pacific Northwest prairies and oak woodlands: challenges, successes, and
future directions. Northwest Science 85:317–328.
Hartnett DC, Wilson GWT (2002) The role of mycorrhizas in plant community structure
and dynamics: lessons from grasslands. Plant and Soil 244:319.
 

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Haskins K, Pence V (2012) Transitioning plants to new environments: beneficial
applications of soil microbes. Pages 89–107 in J. Maschinski, K. Haskins, and P.
Raven, (eds). Plant reintroduction in a changing climate. Island Press/Center for
Resource Economics, Washington, D.C.
Hoeksema JD, Chaudhary VB, Gehring CA, Johnson NC, Karst J, et al. (2010) A metaanalysis of context-dependency in plant response to inoculation with mycorrhizal
fungi. Ecology Letters 13:394–407.
IJdo M, Cranenbrouck S, Declerck S (2011) Methods for large-scale production of AM
fungi: past, present, and future. Mycorrhiza 21:1–16.
Klironomos JN (2003) Variation in plant response to native and exotic arbuscular
mycorrhizal fungi. Ecology 84:2292–2301.
Koide RT, Li M (1989) Appropriate controls for vesicular–arbuscular mycorrhiza
research. New Phytologist 111:35–44.
Kurth F, Zeitler K, Feldhahn L, Neu TR, Weber T, et al. (2013) Detection and
quantification of a mycorrhization helper bacterium and a mycorrhizal fungus in
plant-soil microcosms at different levels of complexity. BMC Microbiology
13:205.
Li A-R, Smith SE, Smith FA, Guan K-Y (2012) Inoculation with arbuscular mycorrhizal
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Richter BS, Stutz JC (2002) Mycorrhizal inoculation of big sacaton: implications for
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with field soil and commercial inoculum for six native montane species and
Bromus tectorum. Restoration Ecology 15:44–52.
Schwartz MW, Hoeksema JD, Gehring CA, Johnson NC, Klironomos JN, et al. (2006)
The promise and the potential consequences of the global transport of mycorrhizal
fungal inoculum. Ecology Letters 9:501–515.
Smith SE, Read DJ (2008) Mycorrhizal symbiosis. Academic Press, Amsterdam, The
Netherlands, Boston, Massachusetts.
Smith SE, Smith FA (2012) Fresh perspectives on the roles of arbuscular mycorrhizal
fungi in plant nutrition and growth. Mycologia 104:1–13.
 
van der Heijden MGA (2004) Arbuscular mycorrhizal fungi as support systems for
seedling establishment in grassland: symbiotic support systems. Ecology Letters
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van der Heijden MGA, Horton TR (2009) Socialism in soil? The importance of
mycorrhizal fungal networks for facilitation in natural ecosystems. Journal of
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Vogelsang KM, Bever JD (2009) Mycorrhizal densities decline in association with
nonnative plants and contribute to plant invasion. Ecology 90:399–407.
 
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Wilson GWT, Williamson MM (2008) Topsin-M: the new benomyl for mycorrhizalsuppression experiments. Mycologia 100:548–554.
Zubek S, Turnau K, Tsimilli-Michael M, Strasser RJ (2009) Response of endangered
plant species to inoculation with arbuscular mycorrhizal fungi and soil bacteria.
Mycorrhiza 19:113–123.

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SUPPLEMENTAL FIGURES AND TABLES


 
Supplemental Figure 1. Perennial Pacific Northwest Willamette Valley–Puget Trough–Georgia Basin
species studied. From left to right, first row: Aquilegia formosa, Balsamorhiza deltoidea, Castilleja
levisecta, Dodecatheon hendersonii, Dodecatheon pulchellum. Second row: Festuca roemeri, Gaillardia
aristata, Micranthes integrifolia, Ranunculus occidentalis, Silene douglasii. Photographs used with
permission of Rod Gilbert.


 
Species
 
Aquilegia
 formosa
 
Balsamorrhiza
 deltoidea
 
Castilleja
 levisecta
 
Dodecatheon
 hendersonii
 
Dodecatheon
 pulchellum
 
Festuca
 roemeri*
 
Gaillardia
 aristata
 
Micranthes
 integrifolia*
 
Ranunculus
 occidentalis
 
Silene
 douglasii
 

Imbibition
 
Stratification
  Imbibition
  Stratification
 
 
time
 (hours)
  time
 (days)
 
date
 
dates
 
11/30/13
 
12/1/13–1/30/14
 
24
 
60
 
12/30/13–1/30/14
 
8
 
30
  12/30/13
 
1/14/13
 
1/14/14–1/30/14
 
12
 
15
 
1/14/13
 
1/14/14–1/30/14
 
24
 
15
 
11/30/13
 
12/1/13–1/30/14
 
24
 
60
 
0
 
0
 
0
 
0
 
10/30/13
 
10/30/2013–1/30/14
 
8
 
90
 
12/30/13–1/30/14
 
12
 
30
  12/30/13
 
1/29/13
 

 0
 
12
 
0
 
12/1/13–1/30/14
 
24
 
60
  11/30/13
 

Supplemental Table 1. Seeds underwent imbibition (soaking) and stratification (dark storage at 3° C)
based on previously established protocols (Krock et al. in review). *Micranthes integrifolia and
Festuca roemeri were added to this study without established imbibition or stratification times.

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Potting soil microbial wash

Restoration prairie microbial wash

High-quality prairie microbial wash

Native AMF

General AMF

Fungicide

Control

NANANAHIHIHIFERO BADE CALE

GEGEGEHIHIHIFERO BADE CALE

FUFUFUHIHIHIFERO BADE CALE

COCOCOHIHIHIFERO BADE CALE

NANANAHIHIHIDOHE DOPU GAAR

GEGEGEHIHIHIDOHE DOPU GAAR

FUFUFUHIHIHIDOHE DOPU GAAR

COCOCOHIHIHIDOHE DOPU GAAR

NAHIMIIN

GEHIMIIN

FUHIMIIN

COHIMIIN

NANAHIHIRAOC SIDO

GEGEHIHIRAOC SIDO

FUFUHIHIRAOC SIDO

COCOHIHIRAOC SIDO

NANANAREREREFERO BADE CALE

GEGEGEREREREFERO BADE CALE

FUFUFUREREREFERO BADE CALE

COCOCOREREREFERO BADE CALE

NANANAREREREDOHE DOPU GAAR

GEGEGEREREREDOHE DOPU GAAR

FUFUFUREREREDOHE DOPU GAAR

COCOCOREREREDOHE DOPU GAAR

NAREMIIN

GEREMIIN

FUREMIIN

COREMIIN

NANARERERAOC SIDO

GEGERERERAOC SIDO

FUFURERERAOC SIDO

COCORERERAOC SIDO

NANANAPOPOPOFERO BADE CALE

GEGEGEPOPOPOFERO BADE CALE

FUFUFUPOPOPOFERO BADE CALE

COCOCOPOPOPOFERO BADE CALE

NANANAPOPOPODOHE DOPU GAAR

GEGEGEPOPOPODOHE DOPU GAAR

FUFUFUPOPOPODOHE DOPU GAAR

COCOCOPOPOPODOHE DOPU GAAR

NAPOMIIN

GEPOMIIN

FUPOMIIN

COPOMIIN

NANAPOPORAOC SIDO

Mycorrhizal Treatments:
NA: native inoculant
GE: general inoculant
FU: fungicide, no inoculant
CO: no inoculant

GEGEPOPORAOC SIDO

Microbial Treatments:
HI: high-quality intact prairie
RE: restoration prairie, likely
out-planting site
PO: potting soil (Sunshine #2)

FUFUPOPORAOC SIDO

Plant Species:
BADE: Balsamorhiza deltoidea
CALE: Castilleja levisecta
DOHE: Dodecatheon hendersonii
DOPU: Dodecatheon pulchellum
FERO: Festuca roemeri

COCOPOPORAOC SIDO

GAAR: Gailardia aristata
MIIN: Micranthes integrifolia
RAOC: Ranunculus occidentalis
SIDO: Silene douglasii

Supplemental Figure 2. Treatments were arranged for each tray to contain one mycorrhizal
treatment, one microbial treatment, and three species. Rectangles represent one tray divided into
three sections, each containing 32 plugs of three species, with 96 plugs per tray.

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Restoration and Its Discontents:
A Critical Analysis of Ethics and Practice in Applied Ecology
INTRODUCTION
Studying the effects of mycorrhizal and microbial organisms on native plants, grown for
outplanting to rare prairie sites, is heavily rooted in the fields of conservation biology and
restoration ecology, as are the values and assumptions behind their practice. Restoration
ecology is a fairly new and highly interdisciplinary area of study with still-developing
philosophies and praxis. Looking critically at some of the problems with both the
underlying philosophies and evolving methods of restoration is especially important as
the goal of research in this field is usually less to inform basic science than to affect
practice. And the stakes are huge—restoration could, as E. O. Wilson believes, be “ . . .
the means to end the great extinction spasm. The next century will . . . be the era of
restoration in ecology” (Wilson 1992). Unfortunately, practitioners could instead be
creating large new problems, encouraging systems that require constant human
intervention, undermining conservation efforts, or simply wasting resources.
The term “restoration” has an inherently positive connotation; it makes us think of
broken things repaired, artworks and buildings returned to their original splendor, and
when combined with “ecology,” of the possibility of returning nature to a pre-human
state. The Society for Ecological Restoration defines it as “ . . . the process of assisting
the recovery of an ecosystem that has been degraded, damaged, or destroyed,” (Clewell et
al. 2004 p. 3, Wilson 1992). Yet restoration ecology has also received deserved criticism,
and many think that the seemingly unimpeachable ethical foundations on which it stands
are actually quite shaky.

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AUTHENTICITY
In “Faking Nature,” an early critique, Robert Elliot (1982) compares environmental
restoration to the forgery of great works of art. Elliot argues that whether or not we are
aware of the deception, something of value is lost in the copy, and that our perceived
value of a work of art or piece of nature is tied up not only in the final result, but in our
assumptions about the process that led to the product. In the case of nature, the
implication is that ecosystems that have come into their current state without human
intervention are more authentic and of greater value than landscapes that have not
experienced human alteration.
Eric Katz (1997) builds on Elliot’s ideas in “The Big Lie: Human Restoration of
Nature,” warning of the dangers of believing that humanity has the ability to restore or
repair the natural environment. Like Elliot, Katz makes distinctions between the natural
and artificial, but develops a more complex argument, acknowledging that the distinction
exists on a continuum rather than as a dichotomy, and that “When we thus judge natural
objects, and evaluate them more highly than artifacts, we are focusing on the extent of
their independence from human domination (p. 396).”
Two main responses to Elliot’s and Katz’s arguments have developed within the
fields of environmental ethics and restoration ecology. The first addresses the difficulties
and ethical problems with separating the concepts of natural from artificial, human from
environment, and attaching value to these. Doing so is not only difficult, potentially racist
and classist, but may ultimately hinder the formation of a viable land ethic that leads to
positive environmental change. The second is that Katz and Elliot’s arguments tend to
focus on taking pristine environments, destroying them, and then restoring them, a

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process that Andrew Light (Light 2000) terms “malicious restoration.” Light argues that
this practice is irrelevant to most ecological restoration projects, which work with
landscapes that were altered by humans, and have existed in that altered state, with no
intentional destruction occuring.
HUMANS, NATURE, AND LINES
Elliot’s line of demarcation between natural and artificial, whether it ever existed at all,
has grown fuzzier for environmental thinkers in the past few decades, perhaps receiving
its final blow in the scientific acceptance that we have entered a new geologic era, the
anthropocene, where human-caused climate change is affecting even Antarctica, the one
continent that appears never to have had major human settlements (Crutzen 2006). This
theoretical line has long been debated by philosophers—are the crops and animals we
grow unnatural? Are they then artifice? Katz would likely point out that these processes
involve domination of nature and are therefore unnatural.

Indigenous people
Indigenous people, and their role in shaping environments, are awkardly missing from
Elliot’s and Katz’s arguments. Richard White addresses some of the problems with
dichotomous environmentalist thinking eloquently:
We are pious toward Indian peoples, but we don’t take them seriously; we don’t
credit them with the capacity to make changes. Whites regularly grant certain
nonwhites a ‘spiritual’ or ‘traditional’ knowledge that is timeless. It is not
something gained through work or labor; it is not contingent knowledge in a
contingent world. In North America, whites are the bearers of environmental
original sin, because whites alone are recognized as laboring. But whites are thus
also, by the same token, the only real bearers of history. This is why our flattery
(for it is usually intended to be such) of ‘simpler’ peoples is an act of such
immense condescension. For in a modern world defined by change, whites are
portrayed as the only beings who make a difference (1996, p. 175).

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Evidence of ancient and extensive indigenous environmental manipulation using
hundreds of strategies is becoming increasingly well-established in the scientific
literature. Even areas deep in the Amazon, originally thought to be truly “untouched”
show evidence of vibrant anthropogenic land management (Posey 1985). Overwhelming
evidence from a wide variety of disciplines indicates that Pacific Northwest prairies and
oak savannas were shaped and maintained by the intentional use of fire for the last
several thousand years (Cooper and Suckley 1859, Boyd 1999, Walsh et al. 2010,
Sprenger and Dunwiddie 2011). Admitting the role of ancient humans in ecosystem
creation and maintenance, however, is damaging to Katz and Elliot’s arguments. If nature
is defined as authentic only when it remains in an untouched state, finding thousands of
years of fingerprints in nearly every corner and crevice of our world leads to the earthshattering conclusion that authenticity does not in fact, exist. Accepting that humans have
left marks everywhere, but belittling the importance of indigenous fingerprints is both
condescending and inaccurate.

Work
It is important to note that only within the last century or so has the view that
human-affected environments may be less valuable than “natural” ones existed to any
significant degree within Western culture (Worster 1994). Rather, wilderness has
historically been seen as a force to be feared and tamed, with the pastoral landscape,
controlled by humans for the creation of food, or the well-ordered flower garden, idyllic.
Landscapes that represented human dominance, food productivity, and labor, were
instead seen as valuable.

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The “leave no trace” ethic of Katz, Elliot, and many environmentalists, is valuable
as a conservation tool, but problematic as a paradigm. This is not only because of its
exclusion of indigenous people and their influence on the land, but the exclusion of all
people and their labor (White 1996). Demarcating land as either for conservation or for
humans may cause us to lose some of our last real connections with natural land. White
argues that humans have long formed connection to nature, and a subsequent
understanding of its value and importance through labor, while modern environmentalists
encourage the formation of this relationship through recreation (1996). He does caution,
however, that physical labor in nature has not historically prevented it from harm. Rather,
enhancing the dichotomous line suggested by philosophers such as Elliot encourages an
unprecedented divorce between nature and work—one that escalates political divides and
alienates people from natural environments.
Our current culture, with its swelling population and economy focused on growth,
requires the setting aside of land for strict conservation. But by focusing on demarcating
and defending the line between humans and nature, rather than acknowledging the blurry
areas that dominate the landscape when one steps back, environmentalists are also
widening divides, and abandoning huge swaths of land that could have been used for
conservation purposes. Failing to see value in land that has been affected by humans and
their labor, the lands that have and will continue to make human existence possible, is a
serious mistake, and one that makes all but the most dedicated primativists into
hypocrites, and some argue, gives environmentalism itself a bad name.

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MALICIOUS RESTORATION
The second rebuttal to Katz and Elliot’s qualms with the “restoration hypothesis” is that
they focus on what Light calls “malicious restorations (Light 2000).” Elliot begins
“Faking Nature” with a hypothetical case in which a company wants to mine a beach for
a particular mineral, which will involve destroying the naturally occurring dunes and
vegetation on the site. However, the company expresses its desire and willingness to
return the beach to its original state after the minerals have been extracted. Light argues
that this sort of malicious restoration should be ethically considered separately from
“benevolent ” restorations, which are carried out to fix harm that has already occurred,
and are not a justification for destructive behavior. There are problems with Light’s
stance, including that unfortunately, malicious restoration is currently a common practice,
that even benevolent restoration can cause harm, and that it can be both difficult and
problematic to draw lines between the two. In fact, both of these types of restoration
utilize much of the same science, and stem from the same logic, the belief that humans
are in fact capable of successfully “fixing” nature (Katz 1997)
Katz, writing later than Elliot, leaves a little bit more possibility for Light’s
benevolent restoration:
I believe, for example, that Exxon should attempt to clean up and restore the
Alaskan waterways and land that was harmed by its corporate negligence. The
point of my argument here is that we must not misunderstand what we humans are
doing when we attempt to restore or repair natural areas. We are not restoring
nature: we are not making it whole and healthy again. Nature restoration is a
compromise; it should not be a basic policy goal. It is a policy that makes the best
of a bad situation; it cleans up our mess. We are putting a piece of furniture over
the stain in the carpet, for it provides a better appearance. As a matter of policy,
however, it would be much more significant to prevent the causes of the stains
(1997, p. 396).

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Both Katz and Elliot fear that an acceptance of the belief that humans can recreate nature
and that what is recreated is of equal value to the original, will lead to thought and policy
that endanger conservation efforts; if we can fix damage that we cause, or effectively
recreate ecosystems in more convenient places, why should we worry about harm?
Malicious restoration has indeed worked its way into current environmental
policy and practice. Legalization officially occurred in the United States in 1993 when
the Clinton administration approved support for the use of “mitigation banks,” and the
US Environmental Protection Agency and Department of the Army issued guidelines for
their establishment and use (Department of the Army, Corps of Engineers Corps of et al.
1995). These policy changes, intented to make conservation under the Endangered
Species Act function within a free-market capitalist system, has validated Elliot and
Katz’s fears (Bayon et al. 2012). Wetlands can be filled and developed if constructed
wetlands are installed elsewhere, forests logged if they are replanted. Development
mitigation allows legal environmental destruction, even where it may otherwise have
been illegal, by allowing companies to purchase environmental credits in the form of
“projects that restor[e], creat[e], enhanc[e], and in exceptional circumstances, preserv[e]”
to mitigate the debit that the original develompent project caused (Department of the
Army, Corps of Engineers Corps of et al. 1995). Restoration organizations, including the
Center for Natural Lands Management, through which this thesis work has been done,
benefit from these sorts of policies, which privatize conservation and make restoration
financially viable.
I believe that we are still at a point in history where, philosophically and
scientifically, re-created nature is not considered to be equally exchangeable with more-

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natural environments, even though legally they may be treated as equivalent (US DOD et
al. 2005). I use the term more-natural here because I don’t agree with drawing a strong
line between humans and nature, but do see a major difference between parcels that have
been subject to restoration activities, and ones that those restoration parcels are emulating.
Ironically, restoration ecologists are likely among those who best understand this
difference.
In the last decade, careful work on the development of locally adapted yet
genetically diverse native seed materials, and the extrapolation of appropriate “seed zones”
within which to source materials has been developing (Bower et al. 2014). Within species,
genotypes are often adapted to very local climatic and ecological variations, and there is
often great diversity even within local genotypes (Rogers & Montalvo 2004). Historically,
the amounts of native seed needed for restoration projects has not been readily available,
and seed materials that are locally maladapted, genetically homogenous, and selected for
by cultivation practices have been introduced to restoration sites, often spreading their
traits through gene flow with wild plants (Bower et al. 2014). As a result, those collecting
native seed from wild populations based on current understanding do their best to avoid
sites where restoration materials may have previously been introduced (US BLM 2014).
Restorationists are quite aware that the practice’s techniques do not yet come close to
actually recreating the authenticity of more-natural sites. The law, unfortunately, does not
seem to have the same scruples.
Will further blurring the lines between restored and more-natural landscapes
lower their value and increase confidence that humans can successfully recreate nature?
Will developing strategies that improve the success of restoration efforts lead to policies

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that promote malicious restoration rather than conservation? Should restorationists who
have benevolent intentions worry that the techniques they develop may be put to use in
unethical ways? Restoration ecology is coming of age, and will need to answer these
Oppenheimer’s dilemma questions that have sprouted from the seeds sown by Elliot and
Katz.

PROBLEMS WITH BENEVOLENT RESTORATION
What will E. O. Wilson’s era of restoration in ecology actually look like? Some, such as
Peter Del Tredici, a plant ecologist who wrote “Neocreationism and the Illusion of
Ecological Restoration,” theorizes that it is most likely to resemble yard work—and a lot
of it:
What’s striking about this so-called restoration process is that it looks an awful lot
like gardening, with its ongoing need for planting and weeding. Call it what you
will, but anyone who has ever worked in the garden knows that planting and
weeding are endless. So the question becomes: Is “landscape restoration” really
just gardening dressed up with jargon to simulate ecology, or is it based on
scientific theories with testable hypotheses? To put it another way: Can we put the
invasive species genie back in the bottle, or are we looking at a future in which
nature itself becomes a cultivated entity (Del Tredici, 2004, p. 2)?
While Richard White might appreciate the implications of a large-scale reintegration of
human labor into nature on some level, it is a problematic prospect. According to Katz,
“A policy of domination subverts both nature and human existence; it denies both the
cultural and natural realization of individual good, human and nonhuman. Liberation
from all forms of domination is thus the chief goal of any ethical or political system
(1997, p. 396).” Restoration is thus solving the problems of human domination through
further domination, a proposal that it is not only theoretically problematic, but has led to

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real problems, such as the introduction of inappropriate “native” plant genotypes (Bower
et al. 2014).
Is restoration, as Del Tredici suggests by mentioning neo-creationism in his title,
an attempt to play god, fighting nature itself? Are we turning ourselves into backwardlooking juvenile deities, still experimenting with how to act, but convinced that this time,
at least, our intentions are good? This comes down to the most difficult social and
environmental question of our time: Can we create change within our existing system, or
will current praxis need to be altered or disassembled on a large scale before a sustainable
future can be reached? Restoration ecology is a child of the current framework, solving
small problems within the limits of the system, while Elliot and Katz advocate a major
shift.

Invasive species and dates
Elliot and Katz are right on some levels, but they are also not ecologists or biologists,
they are philosophers. Invasive species, a product of human travel and globalization, have
become a nearly insurmountable problem in many ecosystems (Pimentel et al. 2005). In
terrestrial environments non-native plants that have traveled with people, escaped from
gardens, or been intentionally introduced often become invasive, dominating the
landscape and outcompeting natives. Ecological checks and balances that kept these
plants under control in their original habitats are absent from their new ones, allowing
them to spread without restrictions, shading out natives, and leaving ecosystems with
greatly reduced diversity, sometimes to the point of monoculture, and even soil microbial
communities and biogeochemical cycles are altered. Plant invasions can lead to the

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extinction not only of native autotrophs, through competition and hybridization, but the
animals that have evolved with native plant species and depend on them for food and
habitat. It is estimated that in the United States there are more than 50,000 foreign species,
that 42% of the species on threatened or endangered species lists are at risk primarily
because of invaders, and that associated economic costs are in excess of $120 billion per
year (Pimentel et al. 2005).
Many restoration ecologists have had to redefine their goals due to the invasive
species problem, and the exorbitant costs (and futility), associated with eradicating them.
Historically, the goal of restoration in the Western hemisphere has been to return
ecosystems to their imagined pre-Columbian state. However, this goal is problematic not
only on practical, but social and philosophical levels.
First, it is very difficult to know with any real certainty what pre-Columbian
ecosystems looked like, partly because landscapes were managed by indigenous people.
Even first explorers’ accounts are not always accurate, for example, in the Pacific
Northwest the spread of disease with population effects more devastating than Europe’s
black plague had reached indigenous people in the 1770s, drastically altering land
management practices long before explorers reached the area (Boyd 1999).
Second, looking at time-scales of thousands of years shows great fluctuation and
variation in vegetation type, driven by changes in climate. In the Pacific Northwest, three
distinct climatic eras have occurred since the last glacial maximum ended 13,900 years
ago, with our modern climate and its associated vegetation only establishing 3,900 years
ago (Whitlock & Bartlein 1997). In fact, the vegetational communities of the early and
late Holocene were driven by a warmer, drier climate than the modern era, and may

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better represent the “natural” vegetation that should occur with global warming.
Attempting to hold ecosystems at a chosen point in time also violates data showing that
the only constant over time has been change, caused by both human and other sources.
Finally, restoring landscapes to imagined pre-Columbian states is impractical. Del
Tredici comments that:
What I find particularly depressing about the ‘native species only’ argument is
that it ends up denying the inevitability of ecological change. Its underlying
assumption is that the plant and animal communities that existed in North
America before the Europeans arrived can and should be preserved. The fact that
this pre-Columbian environment no longer exists—and cannot be recreated—does
not seem to matter (2004 p. 2).
Indigenous people and their land management techniques have largely been replaced with
developed lands, leaving fractured remnant habitats lacking connectivity. Both law and
risk regulate and reduce necessary disturbance regimes, such as fire, and many
ecosystems have been undergoing either a drastically altered or complete lack of
disturbance for centuries. Fencing, hunting, and fragmentation have greatly reduced or
removed keystone animal species, such as large predators. Invasive plants, human
disturbance, pollution, and chemicals have altered soils and microbial communities in
many places. A changing climate will make pre-Columbian restoration even more
difficult.
Herbicide dependency
Herbicide has become one of the major tools currently used by restorationists in the
United States to help leap these hurdles. This is despite the fact that most herbicides were
developed for agriculture, and have not been thoroughly tested in the context of
ecological restoration. Many are known to or suspected of having detrimental effects on

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human health, wildlife, native plants, and soil microbial communities (Wagner & Nelson
2014, Zaller et al. 2014). Katz would argue that using technology such as chemical
herbicides to restore systems is inherently problematic ethically. It is also awkward for a
discipline that espouses to be science-based to depend on practices where insufficient
evidence on long-term effects exists. In fact, evidence showing that even herbicides such
as glyphosate, the most widely used pesticide worldwide and considered one of the most
environmentally friendly can have serious effects on rhizosphere keystone organisms
such as earthworms and arbuscular mycorrhizal fungi, does exist. Recently (Zaller et al.
2014) found that glyphosate significantly reduced root mycorrhization and the biomass of
AMF spores in soil. These detrimental effects occur both through the direct pathway of
herbicide coming into contact with soil, and the indirect pathway of roots via foliage
(Druille et al. 2013).
Restorationists continue to depend heavily on pesticide application despite
scientific evidence and uncertainty because herbicide is one of the few available tools
that allows for more than the smallest of areas to be treated. As humans we generally
depend on visual metrics, and herbicide can provide what we want to see. Unfortunately
many of the side effects, such as soil abiotic changes, microorganism decline, and even
carcinogenic effects are invisible without costly and time-consuming studies and lab
work. Depending on herbicide use without sufficient scientifically based evidence of the
chemicals’ long-term effects may undermine the credibility of restoration ecology within
the scientific community.
An understanding of the harmful effects of pesticides has been part of popular
knowledge since the publication of Rachel Carson’s (1962) Silent Spring and the

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emergence of serious medical problems in Vietnam veterans exposed to Agent Orange.
Support of organic agriculture and skepticism of genetically modified crops and agrochemical companies such as Monsanto, which manufactures glyphosate, have
successfully spread from the environmental fringe to mainstream culture. Restorationists
are likely alienating potential supporters by relying on herbicide use. Indigenous people
may be especially reluctant to support or be involved in restoration projects that use
herbicides, as the practice was introduced by Western culture, has no real counterpart in
traditional management techniques, and could contaminate food and water resources.
This conjecture is based only on hearsay, and research on the opinions of indigenous
people regarding herbicide use in restoration is needed.

NOVEL ECOSYSTEMS
The concept of “novel ecosystems,” which are defined as “[Ecosystems with] species
compositions and relative abundances that have not occurred previously within a given
biome . . . [and] result from biotic response to human-induced abiotic . . . or biotic
elements . . . but do not depend on continued human intervention for their maintenance
(Hobbs et al. 2006 p. 2).” Novel ecosystems can arise either from the degradation of
more-natural ecosystems through human action or plant invasion, or the abandonment of
heavily managed systems, such as agricultural fields (Hobbs et al. 2006). Humans have
created novel ecosystems for millenia, but the number and spatial extent of these systems
has increased rapidly in the modern era. Hobbs et al. (2006) argue that in this historical
context, and because we cannot put Del Tredici’s (2004) “invasive species genie back in
the bottle,” it makes sense to focus value questions on the ability of novel ecosystems to

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provide services of equal value to the ones they are replacing. They argue simultaneously
for:
(1) conserving less impacted places now so they do not change into some new,
possibly less desirable, form; and (2) not wasting precious resources on what may
be a hopeless quest to ‘fix’ those systems for which there is little chance of
recovery back to some pre-existing condition. Rather, we should perhaps accept
them for what they are and what benefits they provide (Hobbs et al. 2004 p. 5).
Elliot and Katz would likely approve of this approach. It supports conservation
and does not allow for malicious restoration. The novel ecosystems theory also leaves
room for change and evolution, and does not attempt to “fix” nature through domination.
These novel ecosystems can instead be seen as a part of humanity, and their current
prevalence as reflective of a culture that values growth and the free hand of the market
over conservation. Instead of denying what we are and what comes with that, this
approach instead asks how we can work with what we have, plan for uncertainty, and
make decisions that will allow novel ecosystems to provide necessary services.

CONCLUSION
Finding solutions to environmental issues, unfortunately, is usually more complex than it
first appears. Even approaches that may initially seem inarguably positive, such as
ecological restoration, are riddled with problems when scrutinized more closely. It is
important, therefore, to know what problematics of logic or ethics commonly come up,
and to develop a set of lenses through which to view and assess proposed environmental
philosophies and methodologies. These should include, but not be limited to, the
following:
1. Is a distinct line drawn between what is human and what is natural?
2. Are indigenous people and their history considered? Involved?

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3. Are the needs of humans, and especially working-class people, considered?
Involved?
4. Is long-term human domination required?
5. Is there the possibility of causing harm?
6.

Are goals realistic and obtainable?

7. Do goals require stasis, or allow for change? Is climate change addressed?
I propose this as a framework for working within our current economic and social
system, but acknowledge that real change to the relationships between humans, nature,
and capital will likely be necessary to create a truly sustainable future. In my lifetime
human-induced climate change has been exposed, gained nearly complete acceptance in
the scientific community, and more recently among those with power in the world’s
governments. However, even with the looming presence of catastrophic warming that
will have devastating effects on every aspect of human life, governments have proven
incapable of agreeing upon or implementing solutions that could drastically slow or halt
climate change (IPCC 2014). Earth has undergone at least five great extinction events
prior to the current anthropogenic one, and life has recovered (Barnosky et al. 2011).
Over geologic time scales, weathering and burial will return CO2 to pre-industrial levels
(Archer 2010). However, it is unknown whether our species will be able to effectively
survive the chaos caused by a warmer climate and other anthropogenic environmental
degradation.
Restoration ecology is a well-intentioned field that has developed some useful
techniques and philosophies, and continues to evolve to address the concerns of its critics.
Yet I do not agree with E. O. Wilson that it will be the means to end the great extinction

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spasm, nor do I hope that the next century will be the “era of restoration in ecology”
unless the discipline’s current ethics and praxis undergo significant change. The good
news is that even as Earth’s population continues to grow and new ecological problems
develop, environmental science and thinking are developing rapidly. New ideas are being
constantly generated and tested, leading to an increase in knowledge that will affect and
improve future practice. Restoration is a major force driving the understanding of both
the potential and limits of applied ecology, and while still young and learning from
mistakes, knowledge gained from today’s experimentation will likely be a part of
solutions in the future.

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