ArticlePDF Available

Abstract

Progress is required in response to how cities can support greater biodiversity. This calls for more research on how landscape designers can actively shape urban ecologies to deliver contextspecific empirical bases for green space intervention decisions. Design experiments offer opportunities for implemented projects within real-world settings to serve as learning sites. This paper explores preliminary ecological outcomes from a multidisciplinary team on whether purposefully engineered native grassland gardens provide more habitat functions for insects than mainstream gardens in the City of Tshwane, South Africa. Six different sites were sampled: two recently installed native grassland garden interventions (young native), two contemporary non-native control gardens (young non-native) on the same premises and of the same ages as the interventions, one remnant of a more pristine native grassland reference area (old native), and one long-established, non-native reference garden (old non-native). Plant and insect diversity were sampled over one year. The short-term findings suggest that higher plant beta diversity (species turnover indicating heterogeneity in a site) supports greater insect richness and evenness in richness. Garden size, age, and connectivity were not clear factors mediating urban habitat enhancement. Based on the preliminary results, the researchers recommend high native grassland species composition and diversity, avoiding individual species dominance, but increasing beta diversity and functional types when selecting garden plants for urban insect biodiversity conservation in grassland biomes.
Citation: Breed, C.A.; Morelli, A.;
Pirk, C.W.W.; Sole, C.L.; Du Toit, M.J.;
Cilliers, S.S. Could Purposefully
Engineered Native Grassland
Gardens Enhance Urban Insect
Biodiversity? Land 2022,11, 1171.
https://doi.org/10.3390/
land11081171
Academic Editor: María Fe Schmitz
Received: 30 April 2022
Accepted: 3 July 2022
Published: 27 July 2022
Publisher’s Note: MDPI stays neutral
with regard to jurisdictional claims in
published maps and institutional affil-
iations.
Copyright: © 2022 by the authors.
Licensee MDPI, Basel, Switzerland.
This article is an open access article
distributed under the terms and
conditions of the Creative Commons
Attribution (CC BY) license (https://
creativecommons.org/licenses/by/
4.0/).
land
Article
Could Purposefully Engineered Native Grassland Gardens
Enhance Urban Insect Biodiversity?
Christina A. Breed 1, * , Agata Morelli 2, Christian W. W. Pirk 2, Catherine L. Sole 2, MariéJ. Du Toit 3
and Sarel S. Cilliers 3
1
Department of Architecture, School of the Built Environment, Faculty of Engineering, Built Environment and
Information Technology, Hatfield Campus, University of Pretoria, Pretoria 0028, South Africa
2
Department of Zoology and Entomology, Faculty of Natural and Agricultural Sciences, University of Pretoria,
Pretoria 0028, South Africa; amorelli028@gmail.com (A.M.); christian.pirk@up.ac.za (C.W.W.P.);
catherine.sole@up.ac.za (C.L.S.)
3Unit for Environmental Sciences and Management, School of Biological Sciences: Botany, Faculty of Natural
and Agricultural Sciences, Potchefstroom Campus, North-West University, Potchefstroom 2531, South Africa;
13062638@nwu.ac.za (M.J.D.T.); sarel.cilliers@nwu.ac.za (S.S.C.)
*Correspondence: ida.breed@up.ac.za
Abstract:
Progress is required in response to how cities can support greater biodiversity. This calls
for more research on how landscape designers can actively shape urban ecologies to deliver context-
specific empirical bases for green space intervention decisions. Design experiments offer opportunities
for implemented projects within real-world settings to serve as learning sites. This paper explores
preliminary ecological outcomes from a multidisciplinary team on whether purposefully engineered
native grassland gardens provide more habitat functions for insects than mainstream gardens in
the City of Tshwane, South Africa. Six different sites were sampled: two recently installed native
grassland garden interventions (young native), two contemporary non-native control gardens (young
non-native) on the same premises and of the same ages as the interventions, one remnant of a more
pristine native grassland reference area (old native), and one long-established, non-native reference
garden (old non-native). Plant and insect diversity were sampled over one year. The short-term
findings suggest that higher plant beta diversity (species turnover indicating heterogeneity in a site)
supports greater insect richness and evenness in richness. Garden size, age, and connectivity were not
clear factors mediating urban habitat enhancement. Based on the preliminary results, the researchers
recommend high native grassland species composition and diversity, avoiding individual species
dominance, but increasing beta diversity and functional types when selecting garden plants for urban
insect biodiversity conservation in grassland biomes.
Keywords:
urban biodiversity conservation; landscape architecture; design; gardens; native species;
grassland plants; insects
1. Introduction
Urbanization and climate change are resulting in an unprecedented loss of biodiver-
sity [
1
]. The Convention on Biological Diversity established firm goals on the integration of
biodiversity into existing and proposed urban development [
2
]. Purposefully planned and
designed urban green infrastructure (UGI) networks can lessen negative urban impacts and
optimize the multifunctionality of limited open spaces [
3
,
4
]. Yet, critical and real progress
is required with regard to how UGI can support greater urban biodiversity [
1
,
5
]. Urban
biodiversity is defined as the variety and richness of living organisms (including genetic
variation) and habitat diversity that is dependent on scale and process, and is found in and
on the edge of human settlements [
1
,
6
]. Although studies exist that demonstrate that cities
can host viable populations of rare or endangered species [
7
], geographic and taxonomic
biases make generalization difficult [
8
]. This is exacerbated by a scarcity of ecological
Land 2022,11, 1171. https://doi.org/10.3390/land11081171 https://www.mdpi.com/journal/land
Land 2022,11, 1171 2 of 25
information specific to geographic areas of a smaller scale [
9
]. Insects are often used as
a biodiversity indicator [
10
,
11
] since they make a large contribution to overall terrestrial
diversity. Together with other invertebrates, they play an important role in ecosystem
processes, such as seed dispersal, predation, and decomposition [12].
Built environment professionals, such as landscape architects, architects, urban de-
signers and planners should purposefully contribute to the provision of ecosystem ser-
vices, wellbeing, and awareness about biodiversity [
3
,
13
]. However, more specific and
locally tested guidelines are needed [
14
,
15
] to enhance the conservation potential of green
spaces [
16
]. Urban environments often create intensively managed, homogenous land-
scapes with distinct intensities and combinations of selective stresses [
17
] and management
preferences [
13
]. This results in the local extinction of many specialist species, which are re-
placed by a few generalist species. This process is referred to as biotic homogenization [
18
].
The restoration of remnant habitats inside urban centers has traditionally held more interest
for biodiversity conservation [
19
]. This paper proposes the construction of small natural
vegetation habitats that could hold potential, but have no guarantee to attract all parts of
the natural biological community [17].
1.1. Literature Review
While purposefully designed (and engineered) green space interventions might in-
crease urban biodiversity, it remains challenging for landscape architects, horticulturalists
and designers to draw any definite conclusions on green space design for conservation
purposes across species. Primary production in urban areas through vegetation preserva-
tion, restoration, reconstruction, and disturbance is directly linked to human decisions and
socioeconomic factors [
20
], yet these become directed by more unintentional determinants
at higher trophic levels [
17
]. We present here a short review of studies that have indicated
vegetation qualities and other factors that influence urban insect biodiversity, deemed
relevant to our study.
In non-urban areas with less human influence, studies have argued that different
vegetation characteristics are responsible for the insect species observed, e.g., plant species
composition [
21
], vegetation height, density, and cover [
22
], plant functional groups [
23
],
or plant beta diversity [
24
]. Habitat patch parameters from traditional conservation biology
have shown the importance of size, structure, and connectivity in urban areas [
25
27
].
General insect habitat guidelines for urban vegetation include biotic complexity, temporal
and spatial heterogeneity (configuration and composition), and unique habitats across
landscape mosaics [
5
,
16
]. While habitat patch age is generally positively correlated with
insect diversity, sealed surfaces and intense management are negatively correlated with
these factors [
28
]. Yet, even small patches of green space have demonstrated value for
certain insects, while rare insect species have shown a preference for native vegetation [
5
].
Insect taxa are therefore considered a useful way of measuring biodiversity [
10
], especially
at a small spatial scale.
Shwartz et al. [
8
] found that 75% of urban biodiversity research was conducted on large
green spaces (over 2 ha), despite studies that show the potential of small patches [
15
,
29
,
30
].
Further research on small habitat patches could shed light on the design and management
requirements of a connecting matrix to support the connectivity and function of habitats [
29
].
Since urban green space is mostly fragmented, this is typically the kind of green space with
which landscape designers work. Small patches thus generally provide fewer resources
and fail to support species diversity through time [
17
]. Fragmentation leads to changes in
behavioral and ecological interactions and processes. It also influences the presence and
relative abundance of species [
17
], yet could be overcome by insects with tolerance and
high colonization potential [28].
The importance of using native plant species in UGI needs greater clarity [
31
,
32
].
The relationship between ecological integrity and native species is mostly presented as
positive and is traditionally favored in restoration projects, but the relationship between
ecosystem services and native status remains contested [
19
,
32
]. Despite the legacy of land-
Land 2022,11, 1171 3 of 25
scape architects such as McHarg [
33
] in applying the theories and principles of ecological
landscape design to urban areas, the relevance of using native species is often debated in
the landscape design industry [
34
,
35
]. Authors question the effectiveness of native species
in providing an optimum habitat for biodiversity, their ability to thrive in urban areas,
and their effectiveness and efficiency in terms of maintenance, risk, and economics [
36
].
These debates, arguably underpinned by value-laden conceptions of nature [
37
] that go
beyond the scope of this paper, are kindled by diverse ecological research narratives and
findings [13].
Recent Global North studies that compared native and non-native plantings in urban
environments in terms of insect habitat show that native plants support greater abun-
dance [
38
] and species richness [
39
]. Yet, both social and ecological studies agree that
non-native plants offer ecological value in urban settings [
38
,
40
], for example, extended
flowering periods [
31
]. A new emphasis on native species has resulted from a greater
uptake of “wilder” urban spaces. These are argued to be more resilient in the provision
of ecosystem services, provide improved habitat, and shape human preference [
39
,
41
].
However, all urban green space needs planning and management to balance the trade-offs
between ecosystem “services and disservices” (so perceived by people).
Built environment designers can influence the functioning of urban ecological aspects
and how people perceive them [
13
,
14
]. Design experiments that are informed by ecological
knowledge and adapted to local conditions can lead to workable local examples that
are monitored and adjusted over time [
3
,
13
,
14
]. Here, we report the first results of the
Biodiversity and Ecosystem Services for Tshwane (BEST) project that investigates landscape
design guidelines to enhance urban biodiversity.
The objectives of the greater project are to study the role of biodiversity and, specifically,
native species in providing ecosystem functions and services, and to consider how city
dwellers’ experience of biodiversity and its services influences their perceptions.
1.2. Local Perspective on Urban Biodiversity
South Africa includes three global biodiversity hotspots and is among the 13 most
megadiverse countries in the world, yet largely makes use of reactive conservation strate-
gies [
42
]. Although the study area of this research is not in a biodiversity hotspot, it is
situated in one of the two most species-rich primary grasslands in the world (with over
2000 plant taxa), occurring in the most urbanized province in South Africa [
43
]. The impact
of urbanization is expected to increase as the country’s urban population of 66.4% in 2018
reaches the projected 79.8% in 2050 [
44
]. The majority of urbanization takes place in the
form of informal settlements, where people are living in poverty and where basic facilities
and services are lacking [
45
]. Apartheid planning resulted in a scarcity of quality urban
green space in marginalized areas, with negative perceptions of often vast, unmaintained,
and unsafe green spaces [
46
,
47
]. Furthermore, a dominance of Eurocentric plant species,
nature forms, and green space activities prevails [47].
As a result, urban green spaces are not always appreciated or perceived as beneficial
in South Africa. Crime is often associated with urban trees and vegetation, yet residents
derive provisioning services such as firewood and shelter [
48
], as well as food and medicine,
and cultural services such as recreation, aesthetic value, and social cohesion from the same
urban trees and vegetation [
49
]. A unique relationship has contextually developed with
urban nature, which has only been partially studied (see refs. [
50
52
]). Efforts are required
from designers to use more local plants in food gardens and to reflect local values and
world views [
47
]. This could foster a greater appreciation of and attachment to the native
environment and native species.
1.3. Objectives
A growing number of studies have considered urban green space conditions and
interventions to improve urban biodiversity conservation (e.g., refs. [
16
,
27
,
28
,
53
,
54
]). This
paper builds on these former efforts with a stronger landscape design and multi- and
Land 2022,11, 1171 4 of 25
interdisciplinary focus. The paper illustrates the first year of findings (January 2019 to
July 2020
) on whether purposefully designed native grassland plant assemblages in ur-
ban green spaces provide habitat functions for biodiversity, as measured through insect
activity. The authors concur with Van Schalkwyk et al. [55] that understanding how plant
diversity reflects insect diversity is important for conservation that is compositionally
and functionally representative. The research specifically aims to answer the following
three questions:
i.
What is the plant species richness, plant species composition, and site heterogeneity
(measured through beta diversity) of the intervention sites (young and native) in
comparison to the control (young and non-native) and reference sites (old and native,
and old and non-native)?
ii.
What differences in ground-dwelling insect species assemblages can be observed
between these different sites?
iii.
What correlations could be drawn between plant species characteristics and observed
insect species diversity?
By answering the above questions, the authors hope to shed light on the question
posed in the title of this paper. Although the empirical work is in its infant stages, the
often short-term duration and constant transformation of urban gardens, synchronized
with rapid urbanization and the advances of climate change, indicate the need to consider
interim results.
2. Method
2.1. Study Area and Study Sites
Gauteng is the economic hub of South Africa—the smallest, most densely populated
province, rich in biodiversity, but pressured by development and rapid urbanization [
43
].
The City of Tshwane metropolitan area lies within a summer rainfall area between the grass-
land and savanna biomes, and contains vegetation comprising high levels of endemism [
56
].
The six areas monitored for this study would once have been covered by the “Moot Plains
Bushveld” [
56
], but also contain grassland vegetation that resembles the Rocky Highveld
Grassland vegetation type, of which only 1.38% is conserved [
57
]. After fynbos, grassland
is the second-most diverse vegetation type in South Africa [
56
]. The temperate grassland
comprises plant communities of grasses and forbs that reappear on an annual basis from
below-ground storage organs [
58
]. Woody plants are rare. The plant biomass is attributed
to the grass component, whereas species richness is attained primarily through the forb
component [58].
Green spaces vary greatly in terms of their insect conservation value according to
their history, size, functional connectivity, heterogeneity and vegetation type, structure,
and volume [
5
]. For similar urban history and functional connectivity, all six gardens
were within a 2 km radius of one another (see Figure 1). Situated on the century-old
campuses of the University of Pretoria, which is located 6 km from the city center, the area
is considered representative of the age and urban development of the city (see Figure 1).
The University’s grounds provide for the ownership, continued access, and management
of the interventions/gardens.
To undertake this research, purposefully engineered green space interventions had to
be implemented. Funding was procured and the interventions’ plants were purposefully
selected. The positions of the “research gardens” (intervention sites) were negotiated as part
of the University’s infrastructure development projects that occurred at the time, namely
the Future Africa Campus and the Javett Art Center. Since perception studies were planned
(not reported in this paper), these premises, which were under development, were selected
as they provide access and use by the general public.
Land 2022,11, 1171 5 of 25
Figure 1.
Location of the intervention, control, and reference areas of the study within Pretoria, City
of Tshwane. Source: Google Earth 2020.
The research team selected two control areas, controlling for the effect of native species
or not, on the same premises and of the same installation ages as the two intervention
sites to account for contextual and environmental factors that cannot be held constant.
Intervention area(s) refer to the young native gardens. Control area(s) refer to the young
non-native gardens. Note that the landscape design consultants determined the plant
species composition and abundance of the controls and the greater infrastructure develop-
ment projects, representing mainstream tendencies in the local green space design industry
without any focus on biodiversity conservation. These included a succulent garden (the
Javett control) and an edible garden (the Future Africa control). Two additional older
reference sites were added to serve as benchmarks for typical local urban habitats: an old
non-native monocot garden and an old native grassland. These older and presumably less
disturbed areas were deemed good reference habitats for insect species encountered in
the city. Habitat patch age is generally positively correlated with insect diversity, while
intense management is negatively correlated [
28
]. The monocotyledon reference garden
contained typical commercial indigenous species that have been popularly used in garden
design in South Africa, while the remnant grassland (the grassland reference) represented
a fairly pristine native grassland area. It was important to select control and reference areas
without trees to enable similar vegetation structures [59] among the gardens.
The patch conditions of the areas are summarized in Table 1. In an attempt to stan-
dardize patch sizes and structures [
26
], the plot sizes were selected to range from 70 to
200 m2
in area, determined by the size of the intervention gardens, with a perimeter length
between 60 and 40 m. The degree of built-up and non-pervious surfaces surrounding the
two intervention sites was not the same, which potentially influenced connectivity [
26
], but
for each intervention site, there was one control and one reference site with similar urban
conditions, but different plant assemblages of similar and older ages (see site details in
Table 1and Figure 1).
Land 2022,11, 1171 6 of 25
Table 1. Sampling areas included in the study with contextual parameters that influence patch conditions.
Plot Name Size (m2) Perimeter (m) Project Start Intervention Surroundings
Moot Plain Grassland
(old and native)
(reference area)
±100 m2sample area
(part of a 5 ha grassland area)
Not applicable for sample area
(buffer area 1100 m perimeter)
Land purchased by the
University of Pretoria in 1938
“Moot Plains Grassland”
Little past disturbance—possibly
grazing. More than 100 different
species previously recorded. Recent
rapid survey recorded 53 species.
Hillcrest Campus: 290 ha, of which
255 ha (92%) is green/pervious.
Large motorway 500 m to the north
and wetland area 1 km to the west.
Monocotyledon Garden
(old and non-native)
(reference area)
±100 m2
(part of 0.25 ha green patch)
Not applicable for sample area
(greater patch 203 m perimeter)
Established as a didactic garden
in the late 1980s
Garden section with 14 mostly
monocotyledon species.
Hatfield Campus: 570 ha, of which
188 ha (32%) is green/pervious.
Larger garden patch includes lawn.
Future Africa Native Garden
(young and native)
(intervention)
198 m260 m October 2018 Garden section with more than
70 different native species planted.
Hillcrest Campus: 290 ha, of which
255 ha (92%) is green/pervious.
Natural rocky outcrop area
500 m to the south.
Future Africa Non-native Garden
(young and non-native)
(control areas)
147 m2+ 37 m249 m + 30 m October 2018
Two garden sections with 12 species.
Same as above.
Javett Native Garden
(young and native)
(intervention)
73 m255 m February 2019 Garden section with more than
40 different species planted.
Hatfield Campus: 570 ha, of which
188 ha (32%) is green/pervious.
Large motorway 3 m to the north,
a dominance of impervious surfaces
.
Javett Non-native Garden
(young and non-native)
(control area)
75 m240 m February 2019 Garden sections with 20 succulent
species planted.
Hatfield Campus: 570 ha, of which
188 ha (32%) is green/pervious.
A dominance of impervious
surfaces, road and trees nearby.
Land 2022,11, 1171 7 of 25
2.2. Plant Species Composition and Species Diversity
The plant species composition for the intervention sites was purposefully selected,
with assistance from native plant growers, to test the species’ urban survival and ability to
enhance biodiversity.
From commercial stock available, several regional evergreen species, flowering forbs
and dwarf shrubs were included for socially aesthetic reasons, since many native grass and
forb species are deciduous or dormant in winter. The plant species were selected to mimic
natural plant compositions, allow for overlapping flowering periods [
60
], and include plant
functional types [
23
] and structural diversity [
59
], which could positively influence insect
diversity and abundance.
The plant species richness, abundance and cover were measured by means of stratified
random quadrat sampling conducted in May and July 2020. Originally planned for April,
but postponed due to the COVID-19 lockdown restrictions, the summer sampling took place
in the beginning of May. Due to warmer urban conditions and irrigation in the gardens,
the sampling was considered representative of the maximum summer growth and cover
for grassland species. The July sampling is representative of the lowest temperatures with
no irrigation in the native gardens, encouraging winter forb, geophyte (including several
woody rootstock species) and grass dormancy and change in species cover dominance.
Circular quadrats with a 1 m radius (3.14 m
2
) were randomly placed in representative
habitat conditions found in each garden patch—with differences in the topography, the
presence of rocks, and variances in species composition. The plant species found in each
quadrat were identified and cover percentage was physically measured. The functional
types (in accordance with [
61
]), permanence, and origin (native range) of each species were
also recorded.
Significant differences between species richness, family richness, and functional types
in the study sites were tested using the Kruskal–Wallis analysis of variance (ANOVA)
and median test, where multiple comparisons of mean ranks for species richness, family
richness, and functional types were generated. In addition, pairwise comparisons were
conducted. The sites were compared in terms of their species composition and cover for
the two sampling periods. Data were square root transformed and compared with the
Bray–Curtis similarity test before being ordinated through non-metric multidimensional
scaling (NMDS) and analyzed through SIMPER to identify the species responsible for
(dis)similarity using Primer v6 [
62
]. The Sørensen coefficient was calculated to determine
and compare the unstructured heterogeneity in species composition [
63
] using the beta
diversity values, which indicate the species turnover between quadrats for each of the sites.
2.3. Insect Diversity and Sampling
The insect diversity in each garden was assessed through pitfall [
16
]. Ground-dwelling
insects were considered representative of the small-scale vegetation refuge function. Sam-
pling was conducted for a week in March, July, and October 2019 to account for seasonal
variation [
64
]. Pitfall traps were inserted in each garden and checked every morning for the
duration of the sampling weeks. They were lined with a standard 50:50 mixture of water
and propylene glycol [
65
]. The traps were arranged at even distances apart and placed to
account for the different plant types and microhabitats present in each garden.
After the specimens from each site had been collected, they were preserved in ethanol
and identified to their lowest classification possible with the aid of specialist taxonomists.
Once identified, species abundance (the total number of individuals found in each garden)
and richness (the total number of different species represented in each garden) [
16
] were
calculated. Diversity indices were also calculated for each garden. The Shannon Diversity
Index accounts for both the richness and evenness of the species present, but is sensitive to
changes in species richness. It assumes that all species are represented in a sample and that
they are randomly sampled [
66
]. The Simpson Diversity Index reflects the species richness
and equitability within communities. It accounts for the proportion of species present in a
sample and gives more weight to common or dominant species [67].
Land 2022,11, 1171 8 of 25
The effect of garden type and age on species abundance and richness, as well as
Shannon and Simpson diversity, was tested using the Kruskal–Wallis ANOVA and median
test, where multiple comparisons of mean ranks for all groups were generated. The effect
of garden age [
28
] on the biodiversity variables was also determined by categorizing each
garden into three age categories: youngest (the Javett gardens, which were about a month
old when the experiment commenced), young (the Future Africa (FA) gardens, which were
less than a year old), and old (the grassland and monocotyledon references, which were
over 20 years of age). The data were processed using the Statistica (Version 64) software
package. The significance level was set at α= 0.05.
2.4. Plant and Insect Sampling Correlations
The differences and/or similarities in plant assemblages and insect diversity and
abundance between the two experimental intervention sites, two controls, and two reference
sites were described using the Spearman correspondence analysis. For the plant and insect
variables, the seasonal totals were used. For the Simpson and Shannon diversity indices,
Sørensen coefficient, and percentage of native plant cover, the seasonal averages for each
garden site were used. The data were processed using the XLSTAT Version 2021.1.1.1088
software package.
3. Results
3.1. Plant Assemblages and Dominant Garden Type Qualities
3.1.1. Plant Species, Family and Functional Type Richness
A total of 93 plant species were recorded from all six sites (list available in
Supplementary
Data). Table 2presents the results from the quadrat sampling in the two seasonal peri-
ods. The gardens are arranged from old (left) to young (right). Sites with native species
dominance are indicated in grey.
The differences in plant species richness (Kruskal–Wallis test: H(5, N= 40) = 25.662,
p< 0.001) (Figure 2a), plant family richness (Kruskal–Wallis test: H(5, N= 40) = 21.481,
p< 0.001) (Figure 2b), and number of plant functional types (Kruskal–Wallis test: H(5,
N= 40) = 22.087, p< 0.001) (Figure 2c) across seasons between the gardens (per quadrat)
were statistically analyzed. The Javett control, FA native garden, grassland reference,
and Javett native garden had comparable species and family richness per quadrat. The
monocotyledon reference and FA control had lower species and family richness. The
FA control showed significantly lower species richness (p= 0.008) and family richness
(p= 0.001) than the FA native garden per quadrat.
Figure 2.
(
a
) The differences in yearly plant species richness, (
b
) plant family richness, and (
c
) plant
functional type richness between sites, measured and compared by quadrat. The bars represent the
median, interquartile range (box), and minimum and maximum (whiskers) of the total. Asterisks
indicate maximum outliers.
Land 2022,11, 1171 9 of 25
Table 2. The mean values of plant variables listed for each site in summer and winter.
Grassland
Reference
(May)
Grassland
Reference
(July)
Monocotyledon
Reference
(May)
Monocotyledon
Reference
(July)
Future Africa
Native Garden
(May)
Future Africa
Native Garden
(July)
Future Africa
Control
(May)
Future Africa
Control
(July)
Javett Native
Garden
(May)
Javett
Native
Garden
(July)
Javett
Control
(May)
Javett
Control
(July)
1. Number of
families 11 12 8 8 18 18 6 6 10 10 7 7
2. Dominant
families
(mean
percentage
coverage)
Poaceae
79%
Asteraceae
9%
Poaceae
78%
Strelitziaceae
33%
Iridaceae
15%
Agapanthaceae
11%
Strelitziaceae
25%
Iridaceae
10%
Asteraceae
23%
Poaceae
19%
Hypoxidaceae
12%
Poaceae
19%
Asteracea
11%
Asphodelaceae
31%
Poaceae 21%
Hypoxidaceae
13%
Asphodelaceae
29%
Poaceae 19%
Lamiaceae
12%
Asphodelaceae
19%
Acanthaceae
15%
Poaceae
13%
Asphodelaceae
19%
Acanthaceae
16%
Poaceae
16%
Aizoaceae
41%
Asphodelaceae
20%
Alliaceae
13%
Aizoaceae
40%
Asphodelaceae
20%
Alliaceae
11%
3. Number of
species 25 27 9 9 32 32 8 8 19 19 14 14
4. Beta
diversity
(mean
dissimilarity
between
quadrats) a
0.60 0.32 0.72
(0.17) y
0.73
(0.20) y0.87 0.85 0.81
(0.43) y
0.87
(0.60) y0.76 0.76 0.27 0.15
5. Dominant
species (mean
percentage
coverage > 8%)
Themeda
triandra 51%
Eragrostis
chloromelas
19%
Themeda
triandra 50%
Eragrostis
chloromelas
19%
Strelitzia juncea
17%
Strelitzia
reginae 17%
Aristea ecklonii
15%
Strelitzia juncea
13%
Strelitzia
reginae 13%
Aristea ecklonii
10%
Haplocarpha
lyrata 19%
Melinis
nerviglumis
18%
Stachys
natalensis var.
natalensis 12%
Melinis
nerviglumis
19%
Haplocarpha
lyrata 9%
Melinis
nerviglumis
21%
Aloiampelos
tenuior 14%
Hypoxis
hemerocallidea
13%
Melinis
nerviglumis
19%
Aloe vera 13%
Melissa
officinalis 12%
Polygala
virgata var.
virgata 10%
Dicliptrea eenii
9%
Dyschoriste
setigera 9%
Aloe merlothii
9%
Carprobrotus
edulis 39%
Nothoscordum
gracile 13%
Aloe
arborescens
(hybrid) 10%
Carprobrotus
edulis 37%
Nothoscordum
gracile 11%
Aloe
arborescens
(hybrid) 10%
6. Plant cover
(%) 100% 78% 95% 56% 89% 58% 89% 80% 87% 71% 100% 96.5%
7. Rock (%) 0% 0% 0% 0% 7% 7% 0% 0% 11% 12% 0% 0%
8. Soil (%) 0% 1% 5% 32% 4% 30% 11% 20% 2% 11% 0% 3.5%
9. Litter (%) 0% 21% 0% 12% 0% 6% 0% 0% 0% 6% 0% 0%
10. Evergreen
plant cover
(%)
4% 0% 80% 52% 39% 30% 34% 32% 55% 50% 100% 96.5%
Land 2022,11, 1171 10 of 25
Table 2. Cont.
Grassland
Reference
(May)
Grassland
Reference
(July)
Monocotyledon
Reference
(May)
Monocotyledon
Reference
(July)
Future Africa
Native Garden
(May)
Future Africa
Native Garden
(July)
Future Africa
Control
(May)
Future Africa
Control
(July)
Javett Native
Garden
(May)
Javett
Native
Garden
(July)
Javett
Control
(May)
Javett
Control
(July)
11. Deciduous
plant cover
(%) 96% 78% 15% 4% 52% 29% 55% 49% 32% 25% 0% 0%
12. Total number
of plant
functional
types
4 4 4 4 6 6 5 5 5 5 2 2
13. Dominant
functional
type (and
species
number
thereof)
Perennial
graminoid
(7)
Perennial
graminoid
(8)
Perennial forb
(6) Perennial forb
(6) Perennial forb
(6)
Perennial
graminoid
(4)
Succulent
(3) Succulent
(3) Succulent
(5)
Perennial
graminoid
(5)
Succulent
(12) Succulent
(12)
14. Native
species
(mean %
contribution
to overall
coverage)
100% 78% 8% 0% 87.75% 60% 43% 37% 87% 71% 14% 14%
15. Indigenous/
non-native
species
(mean %
contribution
to overall
coverage)
0% 0% 59% 40% 1.25% 1% 22% 19% 0% 0% 54% 51%
16. Alien species
(mean %
contribution
to overall
coverage)
0% 0% 18% 11% 0% 0% 24% 24% 0% 0% 23% 21%
ªSørensen coefficient was used to determine the mean dissimilarity between pairs of plots. ySørensen coefficient with outlier plot removed.
Land 2022,11, 1171 11 of 25
The following eight functional types were recorded among the sites: annual/perennial
forbs; geophytes; succulents; annual/perennial graminoids (including grasses and sedges);
dwarf shrubs; and woody shrubs. The Javett, FA control, and grassland reference contained
only an average of two to three functional types per quadrat, compared to the FA and
Javett native gardens and the monocotyledon reference, which contained between four
and five each. The Javett control had significantly fewer functional types than both the
FA (p= 0.004) and Javett (p= 0.008) native gardens. The functional types in the Javett
control were perennial forbs and succulents only. The grassland reference had significantly
fewer functional types than the FA native garden (p= 0.035). The grassland contained
annual/perennial forbs, geophytes, and annual/perennial graminoids, but no dwarf shrubs
or woody shrubs. Notably, only the FA and Javett native gardens had different dominant
functional types between seasons (Table 2).
3.1.2. Plant Species Composition and Cover
The grassland reference, FA, and Javett native gardens have more than 80% native
plant cover, which decreases in winter due to plant dormancy. The FA control garden has
approximately 40% native species cover. The Javett control and monocotyledon references
have a dominant cover of 40 to 50% indigenous species across seasons (see Table 2). The
Javett control fluctuates the least in cover between seasons due to the highest percentage
of evergreen (and succulent) cover, with the monocotyledon reference in second place
(Table 2). The FA and Javett native gardens are the only gardens with different species
cover dominance between the main growing seasons. The FA native garden is the only
garden with different dominant families between full summer cover and extreme winter
cover conditions, alternating between Asteraceae and Poaceae. Noteworthy is that the
grassland reference has 78% grass cover (Poaceae family) throughout the year. The Javett
control is dominated by a single species coverage of 37% and more throughout the year, the
popular succulent creeper Carpobrotus edulis, while the grassland reference is dominated by
a single species coverage of 50% and more throughout the year, namely the grass Themeda
triandra (Table 2).
Tests were conducted to compare the sites in terms of the plant species composition
and cover of all the quadrats in both seasons. With similar patterns in the two tests,
the species cover results, deemed more representative than the species composition, are
discussed here. The stress level was under 0.1 and therefore indicated a good fit. The
NMDS results show that the monocotyledon reference was so compositionally different
to the rest that the remaining sites clustered tightly together (Figure 3a). Note that the
one quadrat of the monocotyledon reference (6.3) was totally dissimilar to the other two
(6.1–6.2), which showed some overlap. By removing the outliers, the monocotyledon
reference (6.1–6.3) and one of the FA control quadrats (2.3 that contained edible exotic
species), the differences between the remaining sites could be investigated (Figure 3b,c).
In both seasons, the grassland (except for one quadrat, 3.2–3.5) and Javett control (5.1–5.2)
showed significant overlap and clustering of quadrats, indicating that they are more similar
in species composition for each site. The proximity of the remaining two quadrats (more
similar) of the FA control (2.1–2.2) suggests some homogeneity in species composition. Both
the FA (1.1–1.4) and Javett (4.1–4.3) native garden showed a widespread arrangement in
quadrats (more dissimilar), indicative of the heterogeneity in species composition between
these gardens.
A SIMPER analysis was conducted where the contributions of individual species to
the separation of the groups could be examined. The analysis revealed that there was
quite a high dissimilarity between all the sites (80% and higher). For the highly clustered
grassland reference, two species contributed more than 80% to a 45% similarity, the grasses
Themeda triandra and Eragrostis chloromelas. In the clustered Javett control, Carpobrotus edulis
contributed 25% to a 72% similarity. The FA control (with the one quadrat removed) had a
40% similarity with two species contributing 99%, namely Hypoxis hemerocallidea and Meli-
nis nerviglumis. The monocotyledon reference had an internal similarity of 20% (reduced by
Land 2022,11, 1171 12 of 25
one very dissimilar quadrat containing Strelitzea reginae and Strelitzea juncea shrubs only);
four species contributed roughly 20% to this similarity, namely Crinum bulbispermum,Aga-
panthus praecox “blue ice” and two exotic weeds, Bidens Pilosa and Pennisetum clandestinum.
The FA and Javett native gardens had an internal similarity of 18% and 14%, respectively.
1
Figure 3.
Non-metric multidimensional scaling of plots for the six gardens across quadrats indicating
(
a
) all quadrats with summer plant cover, illustrating extreme outliers, (
b
) quadrats (outliers removed)
with summer plant cover, and (c) quadrats (outliers removed) with winter plant cover.
3.1.3. Plant Beta Diversity (Species Turnover)
The average species turnover (or beta diversity) between quadrats inside each site
was calculated using the Sørensen coefficient and was reflective of the patterns observed
in the NMDS of plots. The Javett control garden showed very high levels of similarity
in species in summer at 73%. The grassland reference was in second place at 40%. Both
gardens showed even higher similarity in winter at 85% and 68%, respectively (Table 2).
The FA native garden showed the highest beta diversity with a low similarity of 13% and
15% in winter, followed by the FA control at 19% in summer and 13% in winter (note
this position due to one very dissimilar quadrat), the Javett native garden at 24% across
seasons, and the monocotyledon reference at 28% and 27% (note that this position was
due to
one very dissimilar quadrat). To reflect the average beta diversity of the FA control
and monocotyledon reference more accurately, the two extreme outlier plots (2.3 and 6.3)
were removed and beta diversity (Sørensen coefficient) recalculated (Sørensen outliers
removed). The FA control subsequently showed a similarity of species at 57% in summer
and 40% in winter, while the monocotyledon reference showed a similarity of 80% in
summer and 83% in winter (see Table 2).
3.2. Insect Diversity and Abundance
3.2.1. Insect Species Richness
A total of 230 insect taxa were sampled from ten orders across the seasons in the six
areas (see Supplementary Data for species list). The orders samples included Blattodea,
Coleoptera, Dermaptera, Diptera, Hemiptera, Hymenoptera, Lepidoptera, Mantodea,
Orthoptera, and Thysanoptera.
Garden type had a significant effect on species richness in summer (H(
5, N= 42) = 27.78
,
p< 0.001), winter (H(5, N= 42) = 32.11, p= 0.0001), and spring (H(5, N= 42) = 18.08,
p= 0.0028) (see Figure 4, top panel). Species richness was slightly higher in spring
(an average of 58 species in total) than in summer (55 species) and was lowest in win-
ter (43 species). The FA native garden had the highest average and total species richness
across seasons, followed by the monocotyledon and the grassland references. The FA
control and monocotyledon reference, comparatively, maintained slightly higher species
richness numbers in winter, while the other gardens all dropped in richness. The Javett
native garden increased significantly in species richness towards springtime, while the
monocotyledon reference dropped in richness. The grassland reference showed the highest
species richness in summer. The FA native garden showed a significantly higher insect
species richness than the FA control in summer (p= 0.006251), winter (p= 0.011446), and
Land 2022,11, 1171 13 of 25
spring (p= 0.009765), and the Javett native garden showed a significantly higher insect
species richness in summer (p= 0.010574).
Figure 4.
The total insect species richness for each garden (
a
), and the insect species abundance for
each garden (
b
) across seasons. The median, interquartile range (box), and minimum and maximum
(whiskers) of the total. * indicate maximum outliers.
The monocotyledon and grassland references both showed greater species richness
than the Javett native garden in summer (p= 0.045724; p= 0.007363), but not in spring.
3.2.2. Insect Species Abundance
Garden type had a significant effect on total insect abundance in summer (H(5,
N= 42) = 36.57, p< 0.001) (Figure 4a), winter (H(5, N= 42) = 37.00, p< 0.001), and spring
(H(5, N= 42) = 31.37, p< 0.001) (Figure 4b). The highest insect abundance across gardens
was sampled in summer. The monocotyledon reference showed a far greater abundance
than the other gardens, with an average of more than 300 specimens across seasons, and
the greatest numbers in winter. This abundance was significantly greater than the FA
control (p< 0.001 in summer, winter, and spring) and the Javett control (p< 0.001 in
winter, spring, and summer, p= 0.009765). The grassland reference was in second place,
followed by the FA native garden with an average of between 25 and 33 specimens across
seasons. It is, however, important to note that, in both the grassland and monocotyledon
references, of the species encountered, 85% in summer and more than 90% in winter
were Formicidae (ant) species (with as much as 42% in other gardens). The grassland
reference obtained a significantly higher average abundance in summer (119 specimens)
Land 2022,11, 1171 14 of 25
and the Javett native garden obtained a significantly higher average abundance in spring
(65 specimens),
more than
double the abundance they had in the other seasons. The lowest
overall average abundance was obtained by the Javett and FA control (between three and
14 specimens each season).
3.2.3. Shannon Diversity Index (H)
There was an overall effect on insect diversity, as indicated by the Shannon Diversity
Index, in summer (H(5, N= 42) = 29.96, p< 0.001), winter (H(5, N= 42) = 30.93, p< 0.001),
and spring (H(5, N= 42) = 27.92, p< 0.001) (Figure 5a). The FA native garden demonstrated
a significantly greater Shannon Diversity Index across seasons and the monocotyledon
reference consistently demonstrated the lowest Shannon Diversity Index across seasons.
Where the other gardens were quite constant across seasons, the grassland and Javett
controls fluctuated the most, both having a notably lower diversity index in winter. The
FA native garden showed a significantly greater diversity index than the monocotyledon
reference in summer, spring (p= 0.001), and winter (p< 0.001), and the grassland refer-
ence showed a significantly greater diversity index in summer (p= 0.011002) and winter
(p= 0.009382).
Figure 5.
The Shannon Diversity Index for the insect species in each garden (
a
), and the Simpson Diversity
Index for the insect species found in each garden (b) across seasons. * indicate maximum outliers.
Land 2022,11, 1171 15 of 25
3.2.4. Simpson Diversity Index (D)
There was an overall effect on insect diversity, as indicated by the Simpson Diversity
Index, in summer (H(5, N= 42) = 29.21, p< 0.001), winter (H(5, N= 42) = 26.22, p< 0.001),
and spring (H(5, N= 42) = 31.18, p< 0.001) (Figure 5b). The monocotyledon reference
maintained a significantly lower Simpson Diversity Index across seasons. The FA and
Javett native gardens and their controls maintained a fairly consistent diversity index across
seasons. The grassland reference had the second-lowest diversity index in summer and
winter, but rose in spring to overtake the Javett native garden. The Javett native garden
and its control showed significantly greater diversity indices in summer than the grassland
(p= 0.041098; p= 0.016257) and the monocotyledon references (p= 0.001382; p< 0.001).
The FA native garden showed a higher diversity index than the grassland (p= 0.010162,
winter) and monocotyledon (p= 0.001, winter; p< 0.00, spring) controls and the Javett
native garden (p= 0.027561, spring).
3.2.5. Effect of Patch Age
Age had an influence on species richness in summer (H(2, N= 42) = 12.51, p= 0.0019)
and winter (H(2, N= 24) = 8.07, p= 0.0177), but not in spring (H(2, N= 42) = 2.22,
p= 0.3294) (Figure 6a). The older gardens (monocotyledon and grassland references) had
a significantly greater species richness than the youngest gardens (Javett native and its
control) in summer (p= 0.001584) and winter (p= 0.006353), but only slightly greater than
the young gardens (FA native and its control), and not in spring. The young and youngest
gardens showed the greatest richness towards spring and the old gardens in summer.
Figure 6.
The insect species richness in relation to age (
a
), and the insect species abundance in relation
to age (b) across seasons. * indicate maximum outliers.
Land 2022,11, 1171 16 of 25
Garden age impacted on total species abundance in summer (H(2, N= 42) = 28.12,
p< 0.001), winter (H(2, N= 24) = 14.57, p= 0.0007), and spring (H(2, N= 42) = 12.22,
p= 0.0022) (Figure 6b), but was highly influenced by Formicidae species. The young
and youngest gardens obtained a significantly lower abundance than the old gardens
(p< 0.001; p< 0.001) in summer and winter. In spring only, the young gardens obtained a
significantly lower abundance than the old gardens (p= 0.001776). The youngest gardens
had a greater abundance in spring, and the young gardens had a greater abundance in
summer compared to the other seasons. The old gardens had their lowest abundance in
spring and their highest in winter.
Age impacted on overall insect diversity, as indicated by the Shannon Diversity In-
dex in summer, winter, and spring (H(2, N= 42) = 23.33, p< 0.001; H(2, N= 24) = 11.29,
p= 0.0035; H(2, N= 42) = 12.13, p= 0.0023) (Figure 7a). During summer, the old gar-
dens obtained a Shannon Diversity Index that was significantly lower than both the young
(p= 0.001) and youngest (p< 0.001) gardens. In winter (p< 0.001) and spring (
p= 0.001584
), the
old gardens illustrated a significantly lower Shannon Diversity Index than the
young gardens
.
Figure 7.
The Shannon Diversity Index (
a
), and the Simpson Diversity Index (
b
) in relation to the age
of the gardens across seasons. * indicate maximum outliers.
Land 2022,11, 1171 17 of 25
In summer, winter, and spring, age had a significant effect on insect diversity, as shown
by the Simpson Diversity Index (H(2, N= 42) = 28.04, p< 0.001; H(2, N= 24) = 13.76,
p= 0.001; H(2, N= 42) = 20.57, p< 0.001) (Figure 7b). In summer and winter, the older
gardens illustrated a significantly lower Simpson Diversity Index than the young and
youngest gardens. In spring, the young gardens attained a significantly greater diversity
than the youngest (p= 0.030283) and old (p< 0.001) gardens.
3.3. Plant and Insect Sampling Correlations
This study investigated the correlations that could be drawn between plant species
characteristics and observed insect species diversity. The winter correlations are regarded
as being more significant than the summer correlations due to the better seasonal overlap
of the sampling period. Interpretation is subject to the low number of data points (number
of sites). The results of strong and significant correlations will therefore be used as potential
patterns in the data.
For the summer period, no variables indicated strong correlations between plant and
insect data for the sites. Two plant variables showed strong and significant correlations
across seasons, plant functional type and beta diversity (Sørensen coefficient—including
all plots) in summer and winter (rs (4) =
0.9710, p= 0.0028,
0.8827, p= 0.0333, Table 3)
and with beta diversity (Sørensen coefficient—outlier plots removed) in winter (
0.9710,
p= 0.0028). The plant species richness had a strong correlation with the plant family
richness across seasons (p= 0.0333).
For the winter period, there was a strong and significant correlation between plant beta
diversity (Sørensen coefficient—outlier plots removed) and the average Shannon Diversity
Index of insects (p= 0.0167). The plant functional types had a medium-high correlation
with the average Shannon Diversity Index (p= 0.0583).
The strong correlations (not statistically significant) that were found between plant
variables between some of the different study sites (Table 3) confirm that the plants’ species
composition and cover in the interventions, controls, and references were very different, as
indicated in the ordinations. Species-rich gardens were also family-rich, while functional-
type-rich gardens also had high beta diversity (see Table 2). Native range could not be
specifically scrutinized due to the design of the experiment, yet native cover shows a
relationship with beta diversity.
The findings suggest that the high beta diversity (species turnover) between the
quadrats of some of the gardens indicates a heterogeneity in habitats that will influence
a greater number of insect taxa to be present, and the individuals in the community are
distributed more equally among these species.
Land 2022,11, 1171 18 of 25
Table 3. Spearman rho correlations found between plant and insect variables during summer and winter sampling (pvalues in brackets).
Variables
Insect Plant
Abundance Richness Order Ave_
Shannon Index
Ave_
Simpson Index
Ave_Beta
Diversity_All
Ave_Beta Diver-
sity_Removed
Family Richness
( )
Species Richness
( )
Functional Types
( )
Summer
Plants summer
Ave_beta diversity_all 0.3143
(0.5639)
0.0286
(1.0000)
0.3381
(0.4972)
0.3143
(0.5639) 0.1429
(0.8028) - - - - -
Ave_beta diversity_
removed 0.2571
(0.6583)
0.3143
(0.5639)
0.1690
(0.7139)
0.6000
(0.2417)
0.1429
(0.8028) 0.6000
(0.2417) - - - -
Family richness( ) 0.4857
(0.3556) 0.8286
(0.0583)
0.1690
(0.7139) 0.2571
(0.6583)
0.3714
(0.4972)
0.2571
(0.6583)
0.7143
(0.1361) ---
Species richness( ) 0.3143
(0.5639) 0.7714
(0.1028)
0.1690
(0.7139) 0.4857
(0.3556)
0.0857
(0.9194)
0.1429
(0.8028)
0.7714
(0.1028)
0.9429 *
(0.0333) - -
Functional types( ) 0.3237
(0.4972) 0.0883
(0.9194) 0.3482
(0.5639) 0.3825
(0.4972)
0.0883
(0.8028)
0.9710 *
(0.0028)
0.7650
(0.1028) 0.4119
(0.4972) 0.324
(0.564) -
Ave% native cover 0.0286
(1.0000) 0.4286
(0.4194)
0.3381
(0.4972) 0.2571
(0.6583)
0.0857
(0.9194)
0.2571
(0.6583)
0.8286
(0.0583) 0.7143
(0.1361) 0.771
(0.103) 0.441
(0.419)
Winter
Plants winter
Ave_beta diversity_all 0.0857
(0.9194)
0.2571
(0.6583) 0.3395
(0.5639)
0.6571
(0.1750)
0.7143
(0.1361)
Ave_beta diversity_
removed
0.0286
(1.000)
0.4857
(0.3556) 0.1234
(0.9194)
0.9429 *
(0.0167)
0.7714
(0.1028) 0.7714
(0.1028)
Family richness( ) 0.4286
(0.4194) 0.7714
(0.1028) 0.6172
(0.2417) 0.4857
(0.3556) 0.1429
(0.8028) 0.0286
(1.0000)
0.5429
(0.2972)
Species richness( ) 0.1429
(0.8028) 0.5429
(0.2972) 0.4938
(0.3556) 0.5429
(0.2972) 0.2571
(0.6583) 0.1429
(0.8028)
0.4857
(0.3556)
0.9429 *
(0.0333)
Functional types( ) 0.0000
(1.0000) 0.5002
(0.3556)
0.1271
(0.8028) 0.8827
(0.0583) 0.7945
(0.1028)
0.8827 *
(0.0333)
0.9710 *
(0.0028)
0.4119
(0.4972) 0.3237
(0.5639)
Ave% native cover 0.1429
(0.8028) 0.1429
(0.8028)
0.0309
(0.9194) 0.6000
(0.2417) 0.2571
(0.6583)
0.0857
(0.9194)
0.5429
(0.2972) 0.6000
(0.2417) 0.6571
(0.1750) 0.3531
(0.5639)
*Strong correlations (statistically significant but limited due to small data sets).
Land 2022,11, 1171 19 of 25
4. Discussion
The combined findings indicate that the Future Africa native garden (intervention
site) had the highest plant species richness and beta diversity, along with the highest
plant family and functional type richness (Table 2). This garden showed the greatest
consistency in drawing insect diversity throughout the three seasons. It had the highest
insect richness throughout, along with a high evenness according to both the Shannon
and Simpson diversity indices. The grassland reference, with a high native plant species
richness, had high dominance in grass species in both abundance and cover, lower plant
functional type richness, and lower beta diversity (being more homogenous). The grassland
reference still had a general high insect richness and a higher comparative abundance,
yet it demonstrated less evenness in richness for both diversity indices. These findings
suggest that purposefully engineered native grassland gardens could enhance urban insect
diversity.
The exciting prospect for landscape designers is that some of the aspects that drew
insects to the Future Africa native garden, such as varied functional plant types, might also
appeal to people in general. Dwarf shrubs, bulbs, and forbs, for example, form conspic-
uous flowers that are aesthetically pleasing (see Figure 8). Potentially more challenging
aesthetic aspects of grassland plants are that some species, such as geophytes, disappear
underground in winter, while many grass species become dry and brown in color, and
need to be cut short in early spring (see Figure 8, right).
Figure 8.
(
a
c
) The Javett native garden and (
d
f
) the Future Africa native garden in spring (left),
summer (middle), and winter (right).
To respond to the overall question posed in the title and the introduction of this paper,
the authors propose guidelines in each subsection that follows, for grassland plant selection
that enhances urban insect biodiversity conservation. Since the results only account for one
year of monitoring, we specifically discuss them in this regard, resolving ambiguities in the
light of supporting evidence from other studies.
4.1. Plant Beta Diversity (Species Turnover)
Guideline 1: Select a diverse composition (>40) of native plant species.
Guideline 2: Group plant species in very different combinations, in clusters throughout
the garden area.
Land 2022,11, 1171 20 of 25
A definite correlation that could be drawn was that higher beta plant diversity attracted
a greater number of insect species and produced a more equitable distribution among the
insect species present (Table 3). However, this correlation was high when the outlier plant
plots were removed from the beta diversity calculation. The same outlier plot data could
not be identified and removed from the insect data. Regardless, this relationship suggests
that plant species composition and diversity need to be consistent, not merely anecdotal
occurrences, to affect insect richness equitably across taxa.
These findings correspond with recommendations by Kessler et al. [
68
], who advocate
for the measurement of species richness and beta diversity in different landscape elements
(as well as between landscape elements) to advance our understanding of the role of beta
diversity in shaping overall diversity patterns. The importance of plant beta diversity
for dung beetle alpha diversity (measures of species richness) was confirmed by the local
findings of Pryke et al. [
24
] in pine plantations with remnant native grassland and forest
fragments. High grassland beta diversity was associated with high insect alpha diversity
(although larger between-element patterns were also found). Although these inquiries
were at a larger scale, the findings illustrate the importance of high heterogeneity measured
through beta diversity for biodiversity conservation. The findings of this research suggest
that the selected grassland species, quantities, and groupings (creating heterogeneity)
inside the Future Africa native garden were favorable on a small scale and in the short term
for insect urban conservation, exceeding those of remnant native vegetation patches.
4.2. Plant Species Composition and Functional Types
Guideline 3: Select plants from many different families.
Guideline 4: Select plants with a highly diverse (>4) functional structure (geophytes,
forbs, graminoids, succulents, and dwarf or woody shrubs).
Guideline 5: Avoid large areas (>25%) dominated by a single plant species or family
across seasons.
The Future Africa native garden attained a higher insect taxa richness and abundance
than its control garden throughout the year, and the Javett native garden surpassed its con-
trol in winter and spring. These findings, corroborated by those of Botha et al. [
22
], suggest
that the greater diversity of native plants, families, and functional types is more desirable
for larger quantities and varieties of insects, and that the introduction of specific plant
species and their composition may be more valuable than large gardens with less-preferred
plants. Moolman et al. [
69
] found that a large proportion of insects may be host-specific
and therefore likely vegetation-specific because of these relationships. European studies
have shown that the composition of local plant species is the most effective predictor of
insect assemblage composition, even more so than vegetation structure and environmental
conditions [
21
]. Symstad et al. [
23
] found that insect groups were associated with particular
plant assemblages in grasslands, with certain insect orders responding positively to the
increase in specif69 plant functional groups.
Although the grassland reference attracted a higher insect abundance than the Future
Africa native garden, and a high insect richness in general, it fared much poorer in the
Shannon and Simpson diversity indices. A possible reason for this is that the grassland
reference area contains a significantly lower number of functional plant types than the
Future Africa native garden. The grassland reference had a very high dominance of grass
species from the Poaceae family (78% across seasons) and a single grass species dominance
of 50%, which is typical in natural grasslands [
58
]. Comparatively, the Future Africa native
garden had a 19% coverage of grass species, with a high dominance of Asteraceae plants
and other flowering forb families. Several studies confirm the importance of vegetation
heterogeneity (structure and composition) for insect diversity [
5
,
16
]. Notably, the Javett
control garden (succulent garden) also has very few functional types and a single species
dominance of at least 37% across seasons (Table 2). This seems to explain the lower presence
of insects in terms of richness in the Javett control garden.
Land 2022,11, 1171 21 of 25
4.3. Garden Microclimate and Age
Guideline 6: Create small-scale, ecologically distinct landscape features (rocks, dead
wood, ponds, and drier and wetter areas) and varied topography in the garden.
Guideline 7: Even a garden of only a few months old that adheres to guidelines 1 to 6
could attract a great diversity of insects.
In addition to plants in the Future Africa native garden, the landscape heterogeneity
and variance in microclimate may have played a supporting role in creating favorable insect
habitats. Plants significantly affect microclimates by influencing light intensity, temperature,
and humidity [
70
]. For insects, the most important climatic factors are temperature and
water availability [
71
]. The fact that the grassland reference was never irrigated, and
the native gardens were not irrigated in winter, could have influenced the insect species
sampled to some extent [
72
]. For maximizing the numerical and taxonomic diversity of
insects, Clark and Samways [
16
] recommend small-scale, ecologically distinct landscape
features (ecotopes) and varied topography. The ecotopes and topographical variances
apply to the Future Africa and Javett native gardens, but not to the control and reference
areas. The native gardens are on a slight north-facing and west-facing slope, respectively,
and both include clusters of rock.
The Javett native garden is the youngest garden, together with the Javett control
garden. The Javett native garden had the second-highest beta diversity of all the sites (once
outlier plots were removed from the calculations). Noteworthy is that it fared better than
its control in terms of attracting insect richness and abundance in winter and spring. For
the Shannon and Simpson diversity indices, it only surpassed its control in winter, but had
constantly higher indices compared to the monocotyledon reference, and similar or higher
indices than the grassland reference site.
The two older reference sites, the grassland and monocotyledon sites combined,
showed greater abundance and richness in insects than the younger and youngest gardens.
However, during spring, the greater richness pattern was not maintained. This could be
an effect of the rapid establishment period of the younger gardens, with an increase in
species structure and the related microclimate climatic adaptations driving the increase in
insect richness. In general, arthropod species richness declines in cities due to the invasion
and dominance of generalist species [
17
,
18
]. However, Sattler et al. [
28
] found that insect
species richness increased with the age of urban settlements. Some ecological succession
could occur with urban habitat age to create more niches for different species to inhabit, as
well as to allow a higher influx of communities from surrounding areas [
73
]. The results
of this study are therefore interesting with regard to showing the highest taxa richness in
young gardens in spring. Since the native gardens and their non-native counterparts are
extremely novel, with insect sampling commencing when the youngest gardens were a
month old, the effect of patch age will need further monitoring.
The abundance of insects was by far the greatest in the older monocotyledon and
grassland reference gardens, consisting mostly of ant species. Ants are regarded as being
highly competitive due to their invasion ecology [
74
] and because they have broad eco-
logical niches [
75
]. The age of the monocotyledon and grassland reference sites seems to
present plausible explanations for the upsurge in ant abundance.
4.4. Garden Size and Connectivity
Guideline 8: Even a small garden (
±
200 m
2
) that adheres to guidelines 1 to 6 could
attract a great diversity of insects.
Although patch sizes were standardized for sampling, the Future Africa native garden
is the largest recently planted patch. Garden size could potentially play a role in insect
diversity because the larger the garden, the less disturbed the refugia, with potential for
a greater habitat and microclimate effect [
30
]. The large percentage of surrounding green
space at Future Africa, which includes manicured and semi-pristine areas (see Table 1),
could further increase the richness of regional insect species, where impervious surfaces
would be unfavorable [5,26,28].
Land 2022,11, 1171 22 of 25
However, size and green space proportion do not explain the insect richness and
diversity found. Both the grassland reference area and the Future Africa control garden are
on the same premises as the Future Africa native garden (only a couple of meters away from
each other), denoting similar meso-greenspace proportions (see Table 1). The Future Africa
control fared poorer in some seasons in attracting both insect richness and abundance than
both these gardens. The ratio of impervious surface to green space, which negatively affects
insects [
26
,
28
], is less favorable around the Javett native and control gardens (youngest),
which are also half the size of the Future Africa native garden (Table 1). This could have
contributed to the poorer performance of the youngest gardens in terms of insect richness
and abundance. However, habitat size and connectivity do not provide explanations for
the insects that were sampled in this study.
5. Conclusions
This paper specifically aimed to verify the value of introducing purposefully designed
native grassland gardens in urban green areas to support a greater number of insects.
Comparisons between the different sites (each with several sampling plots) in terms of
plant species composition, plant beta diversity, and insect diversity indicated clear patterns.
Although Spearman correlations between some of the plant and insect data were strong
and significant, they only indicate potential patterns due to the small number of sites. These
correlation results are, however, corroborated by other studies that propose the significance
of high plant beta diversity to maintain insect species richness and evenness in richness.
Furthermore, the results confirm that small intervention gardens that are purposefully
planned and designed could rapidly transform urban conditions and provide opportunities
for improved habitats, even compared to remnant native vegetation patches.
For cities in grassland biomes, we propose guidelines for plant selection for urban
insect conservation (for detail, see the Section 4): select native grassland species with a
diverse composition; create a high difference in plant species used in clusters between
different parts of the garden; select species from numerous families; select several different
plant functional types to provide a varied structural strata; avoid individual species and
family dominance; and create varied microclimates by manipulating the topography,
natural features, and levels of humidity/wetness. If the above guidelines are applied, even
small young gardens could make a significant contribution. It is further recommended to
consider principles from traditional restoration ecology in conjunction with the guidelines
provided here. For example, it is important to acknowledge that the inclusion of invader
weeds and exotic plant species in a garden could have a positive influence on plant beta
diversity, but might contradict other ecological restoration goals and ambitions.
The authors believe that this study has created a precedent that must be replicated in
other cities worldwide as a combined effort to assist landscape designers to enhance urban
biodiversity in their local geographical context. The complexity of urban environments
and the many factors that could influence the results of this study illustrate the importance
of other types of interventions, longer-term monitoring, and the comparison of multiple
results to further develop guidelines for more sustainable urban landscapes that enhance
biodiversity conservation. In our study, we focused on ground-dwelling insects only, while
the importance of insect pollinators justifies studies that include a wider variety of insect
sampling techniques.
The ability of native gardens to cope with urban conditions in the long term and
their potential to serve as carbon offsets are aspects that need further research. In the
experience of the authors, the intervention (native grassland) gardens need less watering
and fertilizing than mainstream gardens, but they are not fully self-sustainable. Special
care must be taken to keep invader plant species and weeds under control, while parts
of the garden are bare towards winter and spring. These qualities might create negative
perceptions of native gardens among some people.
Since all gardens need initiative, financing, maintenance, and care, the authors recom-
mend the combination of current ecological studies with social perception studies. The
Land 2022,11, 1171 23 of 25
intervention gardens provide local alternatives to mainstream green space practices and
can provide several ecosystem services, of which aesthetics is but one. However, aesthetic
alternatives are important since they can influence the general acceptance and uptake of
native species. The many possible value relations attached to native species could foster hu-
man sentiments of belonging and care for nature. These human sentiments are imperative
for sustainable and long-term biodiversity conservation.
Supplementary Materials:
The following supporting information can be downloaded at: https:
//www.mdpi.com/article/10.3390/land11081171/s1.
Author Contributions:
Conceptualization, C.A.B.; methodology, C.A.B., C.W.W.P., C.L.S. and S.S.C.;
software, C.A.B., A.M., C.W.W.P. and M.J.D.T.; validation, C.A.B., A.M. and C.W.W.P.; formal analysis,
C.A.B., A.M., C.W.W.P. and M.J.D.T.; investigation, C.A.B. and A.M.; resources, C.A.B., C.W.W.P.
and C.L.S.; data curation, C.A.B. and A.M.; writing—original draft preparation, C.A.B. and A.M.;
writing—review and editing, C.A.B., A.M., C.W.W.P. and S.S.C.; visualization, C.A.B. and M.J.D.T.;
supervision, C.W.W.P., C.L.S. and S.S.C.; project administration, C.A.B., C.W.W.P. and C.L.S.; funding
acquisition, C.A.B., C.W.W.P. and C.L.S. All authors have read and agreed to the published version of
the manuscript.
Funding:
National Research Foundation: Incentive funding; University of Pretoria: University
Capacity Development Program; Research Development Program.
Conflicts of Interest: The authors declare no conflict of interest.
References
1.
Cardoso, P.; Barton, P.S.; Birkhofer, K.; Chichorro, F.; Deacon, C.; Fartmann, T.; Fukushima, C.S.; Gaigher, R.; Habel, J.C.; Hallmann,
C.A.; et al. Scientists’ warning to humanity on insect extinctions. Biol. Conserv. 2020,242, 108426. [CrossRef]
2. Müller, N.; Werner, P.; Kelcey, J.G. (Eds.) Urban Biodiversity and Design; Wiley and Blackwell: Oxford, UK, 2010.
3.
Ahern, J.; Cilliers, S.S.; Niemelä, J. The concept of ecosystem services in adaptive urban planning and design: A framework for
supporting innovation. Landsc. Urban Plan. 2014,125, 254–259. [CrossRef]
4.
McEwan, K.; Ferguson, F.J.; Richardson, M.; Cameron, R. The good things in urban nature: A thematic framework for optimising
urban planning for nature connectedness. Landsc. Urban Plan. 2020,194, 103687. [CrossRef]
5.
Samways, M.J.; Barton, P.S.; Birkhofer, K.; Chichorro, F.; Deacon, C.; Fartmann, T.; Fukushima, C.S.; Gaigher, R.; Habel, J.C.;
Hallmann, C.A.; et al. Solutions for humanity on how to conserve insects. Biol. Conserv. 2020,242, 108427. [CrossRef]
6.
Lindenmayer, D.B.; Fischer, J. Habitat Fragmentation and Landscape Change: An Ecological and Conservation Synthesis; Island Press:
Washington, DC, USA, 2006.
7.
Derby Lewis, A.; Bouman, M.J.; Winter, A.M.; Hasle, E.A.; Stotz, D.F.; Johnston, M.K.; Klinger, K.R.; Rosenthal, A.; Czarnecki, C.A.
Does nature need cities? Pollinators reveal a role for cities in wildlife conservation. Front. Ecol. Evol. 2019,7, 1–8. [CrossRef]
8.
Shwartz, A.; Turbé, A.; Julliard, R.; Simon, L.; Prévot, A.C. Outstanding challenges for urban conservation research and action.
Glob. Environ. Chang. 2014,28, 39–49. [CrossRef]
9.
McDonnell, M.J.; Hahs, A.K. The future of urban biodiversity research: Moving beyond the ‘low-hanging fruit’. Urban Ecosyst.
2013,16, 397–409. [CrossRef]
10.
McGeoch, M.A. The selection, testing and application of terrestrial insects as bioindi-cators. Biol. Rev. Camb. Philos. Soc.
1998
,73,
181–201. [CrossRef]
11.
Helden, A.J.; Stamp, G.C.; Leather, S.R. Urban biodiversity: Comparison of insect assemblages on native and non-native trees.
Urban Ecosyst. 2012,15, 611–624. [CrossRef]
12. Weisser, W.W.; Siemann, E. Insects and Ecosystem Function; Springer: Berlin/Heidelberg, Germany, 2007.
13.
Aronson, M.F.; Lepczyk, C.A.; Evans, K.L.; Goddard, M.A.; Lerman, S.B.; MacIvor, J.S.; Nilon, C.H.; Vargo, T. Biodiversity in the
city: Key challenges for urban green space management. Front. Ecol. Environ. 2017,15, 189–196. [CrossRef]
14.
Felson, A.J.; Bradford, M.A.; Terway, T.M. Promoting earth stewardship through urban design experiments. Front. Ecol. Environ.
2013,11, 362–367. [CrossRef]
15.
Brunbjerg, A.K.; Hale, J.D.; Bates, A.J.; Fowler, R.E.; Rosenfeld, E.J.; Sadler, J.P. Can patterns of urban biodiversity be predicted
using simple measures of green infrastructure? Urban For. Urban Green. 2018,32, 143–153. [CrossRef]
16.
Clark, T.E.; Samways, M.J. Sampling arthropod diversity for urban ecological landscaping in a species-rich southern hemisphere
botanic garden. J. Insect Conserv. 1997,1, 221–234. [CrossRef]
17.
Faeth, S.H.; Bang, C.; Saari, S. Urban biodiversity: Patterns and mechanisms. Ann. N. Y. Acad. Sci.
2011
,1223, 69–81. [CrossRef]
[PubMed]
18.
McKinney, M.L.; Lockwood, J.L. Biotic homogenization: A few winners replacing many losers in the next mass extinction.
Trends Ecol. Evol. 1999,14, 450–453. [CrossRef]
Land 2022,11, 1171 24 of 25
19. Standish, R.J.; Hobbs, R.J.; Miller, J.R. Improving city life: Options for ecological restoration in urban landscapes and how these
might influence interactions between people and nature. Landsc. Ecol. 2013,28, 1213–1221. [CrossRef]
20.
Millennium Ecosystem Assessment (MEA). Ecosystems and Human Wellbeing, Synthesis; Island Press: Washington, DC, USA, 2005.
21.
Knuff, A.K.; Staab, M.; Frey, J.; Helbach, J.; Klein, A.-M. Plant composition, not richness, drives occurrence of specialist herbivores.
Ecol. Entomol. 2019,44, 833–843. [CrossRef]
22.
Botha, M.; Siebert, S.J.; Van Den Berg, J. Do arthropod assemblages fit the grassland and savanna biomes of South Africa?
S. Afr. J. Sci. 2016,112, 9–10. [CrossRef]
23.
Symstad, A.J.; Siemann, E.; Haarstad, J. An experimental test of the effect of plant functional group diversity on arthropod
diversity. Oikos 2000,89, 243–253. [CrossRef]
24.
Pryke, J.S.; Roets, F.; Samways, M.J. Importance of habitat heterogeneity in remnant patches for conserving dung beetles.
Biodivers. Conserv. 2013,22, 2857–2873. [CrossRef]
25.
Collinge, S.K. Effects of grassland fragmentation on insect species loss, colonization, and movement patterns. Ecology
2000
,81,
2211–2226. [CrossRef]
26.
Beninde, J.; Veith, M.; Hochkirch, A. Biodiversity in cities needs space: A meta-analysis of factors determining intra-urban
biodiversity variation. Ecol. Lett. 2015,18, 581–592. [CrossRef] [PubMed]
27.
Norton, B.A.; Evans, K.L.; Warren, P.H. Urban biodiversity and landscape ecology: Patterns, processes and planning. Curr. Landsc.
Ecol. Rep. 2016,1, 178–192. [CrossRef]
28.
Sattler, T.; Duelli, P.; Obrist, M.K.; Arlettaz, R.; Moretti, M. Response of arthropod species richness and functional groups to urban
habitat structure and management. Landsc. Ecol. 2010,25, 941–954. [CrossRef]
29.
Goddard, M.A.; Dougill, A.J.; Benton, T.G. Scaling up from gardens: Biodiversity conservation in urban environments.
Trends Ecol. Evol. 2010,25, 90–98. [CrossRef]
30.
Jaganmohan, M.; Vailshery, L.S.; Nagendra, H. Patterns of insect abundance and distribution in urban domestic gardens in
Bangalore, India. Diversity 2013,5, 767–778. [CrossRef]
31.
Salisbury, A.; Armitage, J.; Bostock, H.; Perry, J.; Tatchell, M.; Thompson, K. Enhancing gardens as habitats for flower-visiting
aerial insects (pollinators): Should we plant native or exotic species? J. Appl. Ecol. 2015,52, 1156–1164. [CrossRef]
32.
Conway, T.M.; Almas, A.D.; Coore, D. Ecosystem services, ecological integrity, and native species planting: How to balance these
ideas in urban forest management? Urban For. Urban Green. 2019,41, 1–5. [CrossRef]
33. McHarg, I.L. Design with Nature; Wiley: New York, NY, USA, 1969.
34.
Kendle, A.D.; Rose, J.E. The aliens have landed! What are the justifications for ‘native only’ policies in landscape plantings?
Landsc. Urban Plan. 2000,47, 19–31. [CrossRef]
35.
Özgüner, H.; Kendle, A.D.; Bisgrove, R.J. Attitudes of landscape professionals towards naturalistic versus formal urban landscapes
in the UK. Landsc. Urban Plan. 2007,81, 34–45. [CrossRef]
36. Zeunert, J. Challenging assumptions in urban restoration ecology. Landsc. J. 2013,32, 231–242. [CrossRef]
37.
Lennon, M. Moral-material ontologies of nature conservation: Exploring the discord between ecological restoration and novel
ecosystems. Environ. Values 2017,26, 5–29. [CrossRef]
38.
Salisbury, A.; Al-Beidh, S.; Armitage, J.; Bird, S.; Bostock, H.; Platoni, A.; Tatchell, M.; Thompson, K.; Perry, J. Enhancing gardens
as habitats for plant-associated invertebrates: Should we plant native or exotic species? Biodivers. Conserv.
2017
,26, 2657–2673.
[CrossRef]
39.
Tallamy, D.W.; Shropshire, K.J. Ranking lepidopteran use of native versus introduced plants. Conserv. Biol.
2009
,23, 941–947.
[CrossRef]
40.
Hoyle, H.; Hitchmough, J.; Jorgensen, A. Attractive, climate-adapted and sustainable? Public perception of non-native planting
in the designed urban landscape. Landsc. Urban Plan. 2017,164, 49–63. [CrossRef]
41.
Threlfall, C.; Kendal, D. The distinct ecological and social roles that wild spaces play in urban ecosystems. Urban For. Urban Green.
2018,29, 348–356. [CrossRef]
42.
Brooks, T.M.; Mittermeier, R.A.; Da Fonseca, G.A.B.; Gerlach, J.; Hoffmann, M.; Lamoreux, J.F.; Mittermeier, C.; Pilgrim, J.D.;
Rodrigues, A.S.L. Global biodiversity conservation priorities. Science 2006,313, 58–61. [CrossRef]
43.
Pfab, M.F.; Compaan, P.C.; Whittington-Jones, C.A.; Engelbrecht, I.; Dumalisile, L.; Mills, L.; West, S.D.; Muller, P.; Masterson,
G.P.R.; Nevhutalu, L.S.; et al. The Gauteng Conservation Plan: Planning for biodiversity in a rapidly urbanising province.
Bothalia Afr. Biodivers. Conserv. 2017,47, 1–16. [CrossRef]
44.
United Nations. World urbanization prospects: The 2018 Revision. United Nations, Department of Economic and Social
Affairs, Population Division (UN/DESA/PD). 2018. Available online: https://population.un.org/wup/Download/ (accessed on
10 October 2020).
45.
Pauleit, S.; Lindley, S.; Cilliers, S.; Shackleton, C. Urbanisation and ecosystem services in sub-Saharan Africa: Current status and
scenarios. Landsc. Urban Plan. 2018,180, 247–248. [CrossRef]
46.
Breed, C.; Mehrtens, H. Using “live” public sector projects in design teaching to transform urban green infrastructure in South
Africa. Land 2022,11, 45. [CrossRef]
47.
Shackleton, C.M.; Gwedla, N. The legacy effects of colonial and apartheid imprints on urban greening in South Africa: Spaces,
species and suitability. Front. Ecol. Evol. 2021,8, 579813. [CrossRef]
Land 2022,11, 1171 25 of 25
48.
Shackleton, S.; Chinyimba, A.; Hebinck, P.; Shackleton, C.; Kaoma, H. Multiple benefits and values of trees in urban landscapes in
two towns in northern South Africa. Landsc. Urban Plan. 2015,136, 76–86. [CrossRef]
49.
Cilliers, S.; Siebert, S.J.; Du Toit, M.J.; Barthel, S.; Mishra, S.; Cornelius, A.; Davoren, E. Health clinic gardens as nodes of
social-ecological innovation to promote garden ecosystem services in sub-Saharan Africa. Landsc. Urban Plan.
2018
,180, 294–307.
[CrossRef]
50.
Cocks, M.; Alexander, J.; Mogano, L.; Vetter, S. Ways of belonging: Meanings of “nature” among Xhosa-speaking township
residents in South Africa. J. Ethnobiol. 2016,36, 820–841. [CrossRef]
51.
Makakavhule, K.; Landman, K. Towards deliberative democracy through democratic governance and design of public spaces in
the South African capital city, Tshwane. Urban Des. Int. 2020,25, 280–292. [CrossRef]
52.
Breed, C.A. Value negotiation and professional self-regulation—environmental concern in the design of the built environment.
Urban For. Urban Green. 2022,74, 127626. [CrossRef]
53.
Norton, B.A.; Thomson, L.J.; Williams, N.S.G.; McDonnell, M.J. The effect of urban ground covers on arthropods: An experiment.
Urban Ecosyst. 2014,17, 77–99. [CrossRef]
54.
Threlfall, C.G.; Mata, L.; Mackie, J.A.; Hahs, A.K.; Stork, N.E.; Williams, N.S.G.; Livesley, S.J. Increasing biodiversity in urban
green spaces through simple vegetation interventions. J. Appl. Ecol. 2017,54, 1874–1883. [CrossRef]
55.
Van Schalkwyk, J.; Pryke, J.S.; Samways, M.J.; Gaigher, R. Congruence between arthropod and plant diversity in a biodiversity
hotspot largely driven by underlying abiotic factors. Ecol. Appl. 2019,29, e01883. [CrossRef]
56.
Mucina, L.; Hoare, D.B.; Lötter, M.C.; Du Preez, P.J.; Rutherford, M.C.; Scott-Shaw, C.R.; Kose, L. Grassland biome. In The
Vegetation of South Africa, Lesotho and Swaziland; Mucina, L., Rutherford, M.C., Eds.; South African National Biodiversity Institute:
Pretoria, South Africa, 2006; Volume 19, pp. 348–437.
57.
Grobler, C.H.; Bredenkamp, G.J.; Brown, L.R. Primary grassland communities of urban open spaces in Gauteng, South Africa.
S. Afr. J. Bot. 2006,72, 367–377. [CrossRef]
58.
Carbutt, C.; Tau, M.; Stephens, A.; Escott, B. The conservation status of temperate grasslands in southern Africa. Grassroots
2011
,
1, 17–23.
59.
Koricheva, J.; Mulder, C.P.; Schmid, B.; Joshi, J.; Huss-Danell, K. Numerical responses of different trophic groups of invertebrates
to manipulations of plant diversity in grasslands. Oecologia 2000,125, 271–282. [CrossRef] [PubMed]
60.
Hunter, M.R. Using ecological theory to guide urban planting design: An adaptation strategy for climate change. Landsc. J.
2011
,
30, 2–11. [CrossRef]
61.
Díaz, S.; Cabido, M. Vive la difference: Plant functional diversity matters to ecosystem processes. Trends Ecol. Evol.
2001
,16,
646–655. [CrossRef]
62.
Clarke, K.R. Non-parametric multivariate analyses of changes in community structure. Austral Ecol.
1993
,18, 117–143. [CrossRef]
63. Vellend, M. Do commonly used indices of β-diversity measure species turnover? J. Veg. Sci. 2001,12, 545–552. [CrossRef]
64.
Buschke, F.; Kemp, M.; Seaman, M.; Louw, S. Intra-annual variation of arthropod-plant interactions and arthropod trophic
structure in an endangered grassland in the Free State province, South Africa. Afr. J. Range Forage Sci.
2011
,28, 57–63. [CrossRef]
65.
Vrdoljak, S.M.; Samways, M.J. Optimizing colored pan traps to survey flower visiting insects. J. Insect Conserv.
2012
,16, 345–354.
[CrossRef]
66.
Spellerberg, I.F.; Fedor, P.J. A tribute to Claude Shannon (1916–2001) and a plea for more rigorous use of species richness, species
diversity and the ‘Shannon-Wiener’ Index. Glob. Ecol. Biogeogr. 2003,12, 177–179. [CrossRef]
67. Simpson, E.H. Measurement of diversity. Nature 1949,163, 688. [CrossRef]
68.
Kessler, M.; Abrahamczyk, S.; Bos, M.; Buchori, D.; Putra, D.D.; Gradstein, S.R.; Höhn, P.; Kluge, J.; Orend, F.; Pitopang, R.; et al.
Alpha and beta diversity of plants and animals along a tropical land-use gradient. Ecol. Appl.
2009
,19, 2142–2156. [CrossRef]
[PubMed]
69.
Moolman, J.; Van den Berg, J.; Conlong, D.; Cugala, D.; Siebert, S.; Le Ru, B. Species diversity and distribution of lepidopteran
stem borers in South Africa and Mozambique. J. Appl. Entomol. 2013,137, 641–720. [CrossRef]
70. Davies-Colley, R.J.; Payne, G.; Van Elswijk, M. Microclimate gradients across a forest edge. N. Z. J. Ecol. 2000,24, 111–121.
71.
Willmer, P.G. Microclimate and the environmental physiology of insects. In Advances in Insect Physiology; Berridge, M.J., Treherne,
J.E., Wigglesworth, V.B., Eds.; Academic Press: Cambridge, MA, USA, 1982.
72.
Cook, W.M.; Faeth, S.H. Irrigation and land-use drive ground arthropod community patterns in an urban desert. Env. Entomol.
2006,35, 1532–1540. [CrossRef]
73. Niemelä, J. Is there a need for a theory of urban ecology? Urban Ecosyst. 1999,3, 57–65. [CrossRef]
74. Rabitsch, W. The hitchhiker’s guide to alien ant invasions. BioControl 2011,56, 551–572. [CrossRef]
75.
Sugihara, G.; Bersier, L.-F.; Southwood, T.R.E.; Primm, S.L.; May, R.M. Predicted correspondence between species abundances
and dendrograms of niche similarities. Proc. Natl. Acad. Sci. USA 2003,100, 5246–5251. [CrossRef]
ResearchGate has not been able to resolve any citations for this publication.
Article
Full-text available
Urban green infrastructure is not acknowledged in the Global South for the critical social and ecological functions it can provide. Contextual design solutions and innovative approaches are urgently needed to transform the status quo. University-local government collaboration could be a way to encourage new thinking, new roles and design skills to develop solutions to these complex problems. This paper presents a case study analysis of such a collaboration. Qualitative research was conducted to establish the degree to which the exposure to real-life projects stimulates postgraduate design students’ transformative learning. The researchers also inquired into the benefits of the collaboration for the municipality. The participants’ reflections were recorded by means of anonymous questionnaires. The findings show that the live project created a municipal setting for seeking alternative solutions in design processes and outcomes. For the students, the project created rich social dynamics and an interplay of familiarity and uncertainty, which aided transformative learning. The students’ deeper learning indicates greater social empathy, reconsidering the role of the profession, greater design process flexibility, and learning and valuing skills across disciplines. The findings hold promise for a more just and sustainable future built environment through collaborations that transform the design professionals involved, the outcomes they pursue, and the processes they follow.
Article
Full-text available
Colonialism is a significant legacy across most aspects of urban form, the nature and distribution of public green spaces, and tree species composition in many cities of the Global South. However, the legacy effects of colonialism on urban green infrastructure and the uses thereof have only recently come under scrutiny. Here we collate information from South Africa on urban greening and interpret it through a colonial and apartheid legacy lens in relation to the distribution and types of urban nature found and their resonance with contemporary needs as an African country. The analysis indicates marked inequalities in public green space distribution and quality between neighborhoods designated for different race groups during the colonial and apartheid periods, which continues to be reproduced by the post-colonial (and post-apartheid) state. Additionally, in the older, former colonial neighborhoods non-native tree species dominate in parks and streets, with most of the species having been introduced during the colonial period. Such colonial introductions have left a burdensome legacy of invasive species that costs billions of Rands annually to keep in check. Lastly, the forms of nature and activities provided in public urban green spaces remains reminiscent of the colonial norm, with little recognition of African worldviews, identity and needs. We conclude in emphasizing the necessity for urban authorities and planners to address these anachronistic legacies through adopting a more inclusive and co-design approach with respect to the extent, location and types of urban nature provided, as well as the types of cultural symbols and activities permitted and promoted.
Article
Full-text available
Many political, economic and social transformations have occurred in South Africa since the first democratic elections in 1994. The country has made significant efforts in trying to establish and rebrand its cities as multiracial and multicultural hubs with democratic public spaces. In an ideal city, public space represents and embodies the ideology of deliberate democracy as postulated by Habermas. However, attempts in South Africa to re(design) public spaces also reflect instances of alienation, conflict and anxiety. This article focusses on the governance and design of public space in the capital of South Africa, the City of Tshwane. The analysis highlights the challenges encapsulated in the governance and design of different types of spaces towards enabling opportunities for deliberate democracy in Tshwane. The paper argues that to address these challenges, urban designers and local authority officials need to focus on both the process and product of urban design, through an emphasis on Spatial Democracy; to readdress governance practices, and Democratic Space; to redirect design practices, both of which have a significant impact on the use, misuse and lack of use of public space.
Article
Full-text available
The fate of humans and insects intertwine, especially through the medium of plants. Global environmental change, including land transformation and contamination, is causing concerning insect diversity loss, articulated in the companion review Scientists' warning to humanity on insect extinctions. Yet, despite a sound philosophical foundation, recognized ethical values, and scientific evidence, globally we are performing poorly at instigating effective insect conservation. As insects are a major component of the tapestry of life, insect conservation would do well to integrate better with overall biodiversity conservation and climate change mitigation. This also involves popularizing insects, especially through use of iconic species, through more media coverage, and more inclusive education. Insect conservationists need to liaise better with decision makers, stakeholders, and land managers, especially at the conceptually familiar scale of the landscape. Enough evidence is now available, and synthesized here, which illustrates that multiple strategies work at local levels towards saving insects. We now need to expand these locally-crafted strategies globally. Tangible actions include ensuring maintenance of biotic complexity, especially through improving temporal and spatial heterogeneity, functional connectivity, and metapopulation dynamics, while maintaining unique habitats, across landscape mosaics, as well as instigating better communication. Key is to have more expansive sustainable agriculture and forestry, improved regulation and prevention of environmental risks, and greater recognition of protected areas alongside agro-ecology in novel landscapes. Future-proofing insect diversity is now critical, with the benefits far reaching, including continued provision of valuable ecosystem services and the conservation of a rich and impressive component of Earth's biodiversity.
Article
Full-text available
Here we build on the manifesto 'World Scientists' Warning to Humanity, issued by the Alliance of World Scientists. As a group of conservation biologists deeply concerned about the decline of insect populations, we here review what we know about the drivers of insect extinctions, their consequences, and how extinctions can negatively impact humanity. We are causing insect extinctions by driving habitat loss, degradation, and fragmentation, use of polluting and harmful substances, the spread of invasive species, global climate change, direct overexploitation, and co-extinction of species dependent on other species. With insect extinctions, we lose much more than species. We lose abundance and biomass of insects, diversity across space and time with consequent homogenization, large parts of the tree of life, unique ecological functions and traits, and fundamental parts of extensive networks of biotic interactions. Such losses lead to the decline of key ecosystem services on which humanity depends. From pollination and decomposition, to being resources for new medicines, habitat quality indication and many others, insects provide essential and irreplaceable services. We appeal for urgent action to close key knowledge gaps and curb insect extinctions. An investment in research
Article
Full-text available
It is well-established that cities need nature for critical ecosystem services—from storing carbon, to reducing temperatures, to mitigating stormwater—and there is growing momentum to seek out strategies for how these services can intersect with urban design and planning efforts. Social scientists and conservation planners increasingly point to urban residents' need to breathe fresh air, encounter the natural world, and have room to play. It is less obvious, perhaps, whether nature needs cities in order to thrive. The evidence from both urban planning and conservation planning is increasingly “yes.” As changes in land use and land cover sweep the planet, cities are becoming important refugia for certain wildlife populations. In recent years, urban planning has embraced the concept of “green infrastructure” as a way to embed green space across metropolitan landscapes to draw on the inherent benefits nature provides to cities, as well as to create habitat for wildlife. We explore this evolving view of cities and nature in the fields of urban and conservation planning. We argue the time is ripe to bring these worlds together, and, using our empirical work, establish that cities matter for monarch butterflies, other pollinators, and at-risk wildlife species.
Article
Built environment design professionals balance competing requirements for human comfort, aesthetics and ecological integrity in their projects. Among them, landscape designers play an important role in optimising the benefits derived from urban green spaces. This study critically considers value dimensions and domains that impinge on professionals’ decision making and environmental values that are applicable to built environment design. It specifically identifies the dominant group-based value domains and environmental values in local practice and examines how landscape designers determine the relative importance of competing values in urban South Africa. The study made use of semi-structured interviews with long-standing figures in the industry. The findings indicate that the political change, the economy and the varied climate and landscape character are contextual influences that strongly shape regional social constructions of urban nature. These constructions emphasise instrumental utility value that juxtapose fundamental values and overshadow less urgent eudemonistic and intrinsic environmental values. In the design process, designers trade off artistic pursuits for instrumental values, with implications for aesthetics and quality. The contextual influences undermine value pluralism. Current trade-offs arguably perpetuate a local legacy of the uneven distribution of quality green spaces. Promisingly, designers show environmental concern and value for sustainability, indigenous plants, utility, quality and safety. The potential for designers to strengthen bonds with urban nature lies in the creation of aesthetic experiences that build on existing local affinities to landscape character and indigenous species. Professional bodies have an important task to assist built environment designers in creating cities that preserve human-nature relationships.
Article
Green interventions which connect people with nature to improve wellbeing are increasingly being applied to tackle the current crisis in mental health. A novel Smartphone app intervention was evaluated amongst adults (n = 228) including (n = 53) adults with common mental health problems, with the aim to improve wellbeing through noticing the good things about urban nature. The app prompted participants once a day over 7 days to write notes about the good things they noticed in urban green spaces. Notes were thematically analysed and ten themes emerged. The three themes with the greatest representation were: i) wonder at encountering wildlife in day-to-day urban settings; ii) appreciation of street trees; and iii) awe at colourful, expansive, dramatic skies and views. Through combining the above themes with the pathways to nature connectedness this paper provides an extended framework of activities to inform activity programming, nature engagement media content, and ‘green health’ interventions. Moreover, the findings have strong implications for optimising city planning, design and management for the wellbeing of both humans and wildlife.
Article
1. How herbivore plant diversity relationships are shaped by the interplay of biotic and abiotic environmental variables is only partly understood. For instance, plant diversity is commonly assumed to determine abundance and richness of associated specialist herbivores. However, this relationship can be altered when environmental variables such as temperature covary with plant diversity. 2. Using gall‐inducing arthropods as focal organisms, biotic and abiotic environmental variables were tested for their relevance to specialist herbivores and their relationship to host plants. In particular, the hypothesis that abundance and richness of gall‐inducing arthropods increase with plant richness was addressed. Additionally, the study asked whether communities of gall‐inducing arthropods match the communities of their host plants. 3. Neither abundance nor species richness of gall‐inducing arthropods was correlated with plant richness or any other of the tested environmental variables. Instead, the number of gall species found per plant decreased with plant richness. This indicates that processes of associational resistance may explain the specialised plant herbivore relationship in our study. 4. Community composition of gall‐inducing arthropods matched host plant communities. In specialised plant herbivore relationships, the presence of obligate host plant species is a prerequisite for the occurrence of its herbivores. 5. It is concluded that the abiotic environment may only play an indirect role in shaping specialist herbivore communities. Instead, the occurrence of specialist herbivore communities might be best explained by plant species composition. Thus, plant species identity should be considered when aiming to understand the processes that shape diversity patterns of specialist herbivores.
Article
Many North American municipalities are developing management plans to guide major investments in their urban forest. These efforts reflect a recent shift in urban forest management, towards ecosystem service provisioning and ecological integrity. The objective of this short communication is to examine how municipal urban forestry plans balance ecosystem services and ecological integrity through a case study of Ontario, Canada. Our review found that all plans emphasize ecosystem service provisioning and ecological integrity, as well as the importance of planting native species, but the majority of plans do not operationalize these ideas. Furthermore, their is tension and uncertainty between managing for ecosystem services and ecological integrity that is neither acknowledged nor addressed. We argue that the tension between ecosystem service provisioning and ecological integrity related to the role of native species can be reduced by recognizing that non-native species are appropriate to achieve both ideas in urban landscapes. A shift away from prioritizing native species would also better align management plans with current practice.