Birds, Bees, and Botany: Measuring Urban Biodiversity After Nature-Based Solutions Implementation

Abstract

Nature-based Solutions (NbS) are increasingly adopted in urban settings to restore ecological functions and enhance biodiversity. This study evaluates the effects of NbS interventions on bird, insect, and plant communities in the Cavalum Valley urban green area, Penafiel (northern Portugal). Over a three-year period, systematic field surveys assessed changes in species richness, abundance, and ecological indicators following actions such as riparian restoration, afforestation, habitat diversification, and invasive species removal. Results revealed a marked increase in bird overall abundance from 538 to 941 individuals and in average pollinator population size from 9.25 to 12.20. Plant diversity also improved, with a rise in native and RELAPE-listed species (5.23%). Functional group analyses underscored the importance of vegetative structure in supporting varied foraging and nesting behaviours. These findings highlight the effectiveness of integrated NbS in enhancing biodiversity and ecological resilience in urban landscapes while reinforcing the need for long-term monitoring to guide adaptive management and conservation planning. Future work could evaluate ecological resilience thresholds and community participation in citizen science monitoring.

1. Introduction

Urban green spaces have become increasingly important as ecological refuges and providers of environmental services due to rapid urbanization and city expansion. These valuable ecosystems provide a range of services, including habitats for diverse fauna, including endangered species. Furthermore, these areas play a critical role in regulating the water cycle, mitigating climate change impacts, providing shade, and reducing the urban heat island effect [1].

Despite their ecological value, urban green areas are increasingly threatened by both direct and indirect pressures. Direct threats include urbanization, land-use changes, and soil sealing, which contribute to habitat fragmentation, disruption of hydrological cycles, and the breakdown of trophic interactions [2]. Indirect pressures stem from pollution caused by urban runoff and nearby agricultural activities, which can degrade water quality and soil health [3]. Additionally, excessive anthropogenic use, such as trampling, recreational overuse, and improper waste disposal, can lead to the progressive degradation of these ecosystems. These stressors disproportionately affect sensitive species, which may be unable to adapt to rapid environmental changes, resulting in biodiversity loss and compromised ecosystem functioning [4].

Urban planners are increasingly adopting habitat units, rather than generic green spaces, to maximize multifunctional outcomes. For example, in Berlin, mixed alder–willow wetland corridors elevated pollinator movement by over 30%, while in Tuscany, inter-rows of cover crops within vineyards bolstered predatory arthropod abundance by nearly 40% [5]. Meanwhile, Mediterranean green-roof retrofits in Barcelona reduced rooftop temperatures by 2 °C and supported diverse native forb assemblages, and targeted flagship bird trails in coastal wetlands have been shown to enhance both avian conservation and community engagement [6].

Ecological monitoring is crucial to ensure these areas’ sustainable development and preservation. By monitoring these ecosystems’ health, city planners can identify areas that require restoration and/or protection, promoting ecological recovery and preventing new environmental pressures. Ecological recovery monitoring also helps identify changes in cultural ecosystem services over time, such as recreation, esthetic value, and opportunities for education and cultural heritage preservation, which is important to recognize progress and identify areas for improvement [7].

Recent reviews also highlight persistent design and monitoring shortfalls. The authors of Castelo, et al. [8] identify four priority research needs and argue that only adaptive, long-term pilot schemes can bridge the gap between theory and practice. Similarly, green-roof and wall systems, when constructed with variable substrate depths and indigenous planting palettes, have been shown to support assemblages of arthropods, bats, and even small passerines, effectively linking core habitat patches across dense urban matrices [9].

Birds are widely recognized as indicators of ecosystem health, given their sensitivity to environmental changes and critical role in maintaining ecological balance within trophic chains [10]. Their wide distribution and visibility make birds easy to observe and count, while birdwatching further enhances community engagement [11]. Consequently, birds provide valuable insights for monitoring ecosystem health, as it responds rapidly to environmental shifts, exhibiting changes in abundance, distribution, and behaviours such as feeding and nesting [12].

Similarly, assessing the richness and distribution of pollinators and insects, together with flora surveys, enables researchers to monitor the presence or absence of these species over time [1]. Observing such population shifts helps reveal the impacts of human activity on ecosystems and identifies areas that require protection or restoration. Furthermore, these species are reliable indicators of ecosystem health across diverse habitats, including forests, grasslands, wetlands, and urban areas [13,14]. Importantly, native flora plays a vital role in sustaining these biological communities by providing key resources for pollinators and insectivorous birds.

Biodiversity monitoring, particularly through field surveys, is a valuable tool for assessing the effectiveness of Nature-based Solutions (NbS) in fostering ecological recovery within urban green areas. Tracking the presence or absence of key species before and after the NbS implementation provides critical insights into ecosystem health and the success of restoration actions [15,16]. These observations also support adaptive management, enabling necessary adjustments to restoration strategies. Different types of NbS may influence animal populations in distinct ways, especially in the context of urban forest management [15]. For instance, green roofs and green walls can enhance bird species richness and abundance by providing additional habitats and food sources such as seeds and insects [17]. In contrast, features like rain gardens and bioswales may not directly increase bird populations but can still provide indirect benefits by improving water quality and increasing vegetation cover [17,18].

Longitudinal comparisons of bird, insect, and flora data, collected before, during, and after the NbS implementation, offer a broader understanding of restoration effectiveness, especially since NbS often operate at the landscape scale [19]. Field surveys are highly adaptable to diverse urban environments, from large parks to fragmented green spaces, and can be tailored to focus on specific taxa or ecosystem functions [20]. Ultimately, biodiversity monitoring generates essential indicators to evaluate the performance of NbS, ecological measure ecological recovery, and ensure the sustained delivery of key ecosystem services over time. Without consistent systematic monitoring, restoration efforts risk underachieving their goals, and the long-term resilience of urban ecosystems remains uncertain [21,22].

This study aimed to evaluate the impact of NbS on biodiversity by using field surveys to assess its effectiveness in restoring riverine habitats and forest ecosystems within an urban setting. We aimed to answer the following questions: (i) to what extent do NbS contribute to the restoration of biodiversity in urban green areas? (ii) How do different types of NbS interventions influence the species richness, abundance, and functional traits of bird, insect (pollinator), and plant communities in an urban context? (iii) Can field-based biodiversity monitoring effectively capture ecological changes over time following the NbS implementation in urban landscapes? (iv) What is the relationship between vegetation structure and specific birds’ functional groups in restored urban habitats? (v) Are there observable trade-offs or synergies in biodiversity outcomes across different taxa (birds, insects, plants) after NbS application?

2. Materials and Methods

2.1. The Study Area

Penafiel city is located in northern Portugal, within the Penafiel municipality (Figure 1), and covers an area of 22.52 km². Its urban development has lacked adequate planning and land management, resulting in various societal challenges, including climate change impacts, flood risk, water pollution, ecosystem degradation, biodiversity loss, and adverse effects on human health—issues that are common to many other European cities [2,23].

The Urban Green Area of the Cavalum Valley (see Figure 2), which is crossed by the Cavalum River, comprises 92 ha and is divided into four homogeneous landscape units. These Urban Green Areas (UGAs) include: the City Park (UGA-A), a landscape garden; a transitional zone (UGA-B) between the Park and a natural oak forest, with a high presence of invasive alien species; the Magikland Park (UGA-C), a thematic park with green areas and private farms; and natural areas (UGA-D), featuring small waterfalls and native forest patches near an industrial settlement.

Biodiversity surveys in the UGAs were conducted at multiple sampling points: two for avifauna in the UGA-A and one in each of the remaining UGAs. The distribution of insect/pollinator survey points followed the same structure as bird surveys, but in duplicate. For vegetation, two points were surveyed in the UGA-B and six across UGAs C and D.

The Cavalum Valley faces several ecological challenges, including poor land management, invasive alien species, and pollution, all exacerbated by population density. Like most European cities, this area is prone to air pollution [25], posing health risks to local residents. Thus, the Cavalum Valley plays a critical role as a green barrier against airborne particles and noise pollution [26]. Water pollution from industrial and agricultural sources also threatens the Cavalum River. To address these challenges, various NbS were implemented, as outlined in

Table 1. NbS-type interventions implemented in the different UGAs (A: City Park; B: Transitional Zone; C: Magikland; D: Natural Area).

NbS-Types	NbS Intervention	Objective	UGA
Ecological Restoration	Recovery and maintenance of riparian vegetation	Reduce water temperature, increase infiltration, reduce riverbank erosion, protect habitats, and increase recreational value, through autochthonous species plantation	A, B and D
	Natural meadows and pastures seeding	Improve carbon sequestration, enhance biodiversity and improve water quality, through seeding of natural meadows.	C
Ecological Engineering	Natural engineering	Creation of habitats for fauna, stabilize riverbanks, consolidate and promote the accumulation of sediments in specific spots with application of braided Salix atrocinerea.	A and D
Forest Landscape Restoration	Afforestation and reforestation with endemic species	Increase tree coverage of endemic species to enhance carbon sequestration; increase endemic biodiversity; control of soil erosion; increase human well-being, by providing areas for recreational activities.	A, B and D
Ecosystem-based management	Increasing habitat heterogeneity	Promote an increase in biodiversity and enhance wildfire resilience.	B and D
	Removal of invasive species	Improving endemic biodiversity and allow native species to prosper.	All
Natural infrastructure	Creation of wetlands	Provide new habitats, water purification through aquatic plants, and increases in cultural ecosystem services, such as esthetic and recreational value or environmental education.	C and D

, to support the ecological restoration of the area. The implementation of NbS started in March 2021 with the first removal of invasive alien species across the UGAs B to D. This clearance was repeated with volunteers at six-month intervals to prevent reinvasion and ensure the sustained recovery of native communities. In October 2021, afforestation and reforestation works began in the UGA-B, followed shortly thereafter in the UGA-C. Seeding of natural meadows was scheduled for early spring 2022, optimizing germination under milder temperatures. Prior to the onset of spring 2022, natural engineering measures were installed along riparian zones to stabilize banks and promote native vegetation. Between late spring and early summer 2022, when hydrological and climatic conditions permitted, small wetland ponds were created. All planned NbS interventions were completed by October 2022, after which ecological monitoring has continued on a seasonal basis.

2.2. Biodiversity Baseline Assessment

In March 2021, prior to the implementation of NbS, a data search was conducted using the Global Biodiversity Information Facility (GBIF), resulting in a baseline species list compiled from citizen science platforms and additional surveys from national, regional, and local databases. This baseline served as a reference to assess species presence or absence before the NbS implementations, establishing a foundation for subsequent biodiversity monitoring.

2.3. Bird Survey Methodology

Bird surveys were conducted on 2 and 25 April 2022 (Year 1), 1 and 25 April 2023 (Year 2), and 9 September 2024 (Year 3). All surveys took place at dawn under similar weather conditions: temperature ranging between 16 °C and 22 °C and low wind.

Visual and sound observations were carried out at five sampling points, each spaced 250 m apart, using binoculars (Leica, Wetzlar, Germany, 8 × 42). At each location, a 20 min observation was conducted within a 50 m radius, with all sightings recorded on field sheets. GPS coordinates were used to geotag each location to ensure consistency across years.

The methodology followed the guidelines of the SPEA (Sociedade Portuguesa para o Estudo das Aves—Portuguese Society for the Study of Birds) for the Common Birds Census in Portugal [27] as well as the protocols by Sauer [28]. When needed, species were identified using the Guide for Birds (Guia de Aves) from the SPEA [27].

Bird species were classified into functional groups based on foraging behaviour, nesting site, and diet type, as described in

Table 2. Functional groups for bird species according to ecological traits (adapted from Mekonen [18]).

Functional Group	Category	Description
Foraging	Ground	Take most food from the ground. Includes species that perch in vegetation and prey from the ground
	Shrubs	Mostly forages in vegetation ≤1 m above ground
	Trees	Mostly forages in vegetation >1 m above ground. Includes species that excavate bark
	Air	Catches their food flying
	All	Forages throughout the range of vertical strata, from ground to the canopy
Nesting	Burrow	Nests in a tunnel in the ground
	Tree Branches	Constructs nest in trees >1.5 m, in branches or flat parts
	Shrubs/Bushes	Nests in shrubs, bushes, or small trees <1 m
	Hollow	Nest in a large or small cavity in a tree
	Artificial	Nests in anthropic cavities, structures, or others
	Other	Not defined or that does not belong in the above definitions
Diet	Frugivore	Mostly raw fruits or succulent fruit-like
	Granivore	Mainly seeds or grains
	Insectivore	Mainly insects
	Nectarivores	Main food item is nectar from flowers
	Herbivores	Eat different parts of plants
	Omnivore	Cannot be differentiated by any type of food
	Scavengers	Feed on carrion
	Carnivores	Hunts other animals

.

2.4. Insects and Pollinators Survey Methodology

Pollinators surveys were conducted on 2 March and 2 April 2022 (year 1), 1 March, 1 April 2023 (year 2), and 9 September 2024 (year 3). At five different points along the Cavalum Valley, a flowering plant was randomly selected, and only insects in direct contact with flowers were considered pollinators. Observations were made near the bird survey points. A total of twenty plants were observed for twenty minutes each, with the observer maintaining a minimum distance of one metre from the plant. All insects present within the observation field were recorded. Most insect species were identified in the field without being captured; however, some individuals, were collected using a standard butterfly net and identified on-site. Taxa were classified to the order level.

2.5. Vegetation Survey Methodology

Vegetation surveys were carried out on 2 April 2022 (year 1), 1 April 2023 (year 2), and 9 September 2024 (year 3), using 20 × 20 m square plots (400 m² each), distributed across eight sites throughout the Cavalum Valley, for a total annual surveyed area of 3200 m². In the UGA-A (City Park), sampling was conducted only once due to the area’s garden characteristics and low expected vegetation change. All vascular plant species rooted within the plots were identified to species level, with their height and coverage abundance classified according to

Table 3. Categories selected and a respective code for flora size and coverage.

Code	Category	Size	Code	Coverage
T	Tree layer	>5 m	5	[76%; 100%]
St	Shrub layer	Between 2 m and 5 m	4	[51%; 75%]
Ss	Shrub layer	Between 0.50 m and 2 m	3	[26%; 50%]
Ht	Herb layer	Between 0.30 m and 1 m	2	[6%; 25%]
Hm	Herb layer	Between 0.10 m and 0.30 m	1	[1%; 5%]
Hs	Herb layer	<0.10 m	r	<1%
ML	Ground/Rocks	<0.05 m

. Rare, Endemic, Localized, Threatened, or Endangered (RELAPE) species were counted and recorded per plot.

2.6. Statistical Analyses

Species richness, abundance, and average population size were calculated for each sampling point, the UGA, and year. In addition, diversity was assessed using the Simpson’s Diversity Index and the Shannon–Wiener Diversity Index.

Species richness was calculated as the total number of species recorded for each sampling point. Mean abundance per sampling point was calculated as follows:

$$A={\displaystyle \frac{{\sum }_{i=1}^N{n}_i}{N}}$$

where $n_i$ is the number of individuals observed at point i, and N is the total number of sampling points.

The Shannon–Wiener Diversity Index ($H^{\prime }$) was calculated as follows:

$$H^{\prime }=-\sum _{j=1}^s{p}_j\ln (p_j)$$

where $p_j={\displaystyle \frac{n_j}{{\sum }_{k=1}^s{n}_k}}$ is the proportional abundance of species j.

Simpson’s diversity index ($D$) was expressed as the inverse of the sum of squared species proportions:

$$D={\displaystyle \frac{1}{{\sum }_{j=1}^s{p}_j^2}}$$

In addition, we quantified dominance using two complementary measures. Simpson’s Index of Dominance ($D^{\prime }$) was calculated as follows:

$$D^{\prime }={\sum }_{i=1}^s{p}_i^2$$

where $s$ is species richness, and $p_i$ the relative abundance of species i. Values range from 0 (perfect evenness) to 1 (complete dominance by a single taxon).

Cumulative Dominance (CD) was calculated by obtaining first the relative abundance of each species as follows:

$$p_i={\displaystyle \frac{n_i}{N}}$$

where $n_i$ is the number of individuals of species $i$, and $N={\sum }_{i=1}^s{n}_i$ is total abundance. Species with $p_i$ ≥ 0.0499 (≈5%) were classified as dominant, and $CD$ was the sum of their relative abundances:

$$CD=\sum _{p_i\ge 0.0499}{p}_i$$

Therefore, $CD$ ranges from 0 (no species exceeds the 5% threshold) to 1 (all individuals belong to dominant species) and is then expressed as a percentage.

3. Results and Discussion

3.1. Temporal Changes in Species Richness and Abundance

The decrease in bird richness observed in 2023 (

Table 4. Richness, average population size and total number of individuals of birds () and insects () species for the Cavalum Valley per year, as well as the Shannon–Wiener Diversity Index, Simpson’s Reciprocal Index, Simpson’s Index of Dominance, and Cumulative Dominance (%).

	Baseline		2022		2023		2024
Richness	54	11	43	74	42	69	43	74
Average population size	5.63	1.18	12.51	9.25	20.17	11.34	21.84	12.20
Abundance	304	13	538	685	857	783	941	903
Shannon-Wiener Diversity Index	3.64	2.35	3.33	3.93	3.27	3.84	3.33	3.85
Simpson’s Reciprocal Index	31.65	9.94	20.97	40.25	19.55	33.80	21.12	34.56
Simpson’s Index of Dominance	0.032	0.101	0.048	0.025	0.051	0.030	0.048	0.029
Cumulative Dominance (%)	50.3	100	46.3	22.4	52.7	30.1	47.6	10.1

) can be attributed to an intervention aimed at removing invasive species, such as Tradescantia fluminensis Vell. and Acacia spp., carried out one week before the field survey by a local NGO (Cavalum—Associação para a Defesa do Ambiente). This activity likely affected biodiversity in the UGA-D, particularly in patch 5, and may also have influenced vegetation structure in spots 6, 7, and 8. The involvement of citizens in removing invasive vegetation might have introduced challenges in accurately identifying plant species [7,29,30,31]. However, during the field survey on 9 September 2024, full biodiversity recovery was evident across all observation points in the valley, as shown in Table 4. Bird richness and abundance had stabilized to levels comparable to 2022. The average bird population size increased from 5.63 at baseline (2021) to 12.51 in 2022, 20.17 in 2023, and 21.84 in 2024. It is important to note that baseline data from 2021 should be interpreted with caution, as they originated from citizen science platforms that may introduce bias or errors due to variability in equipment, observer expertise, and environmental conditions [31]. Nonetheless, these platforms are valuable for fostering environmental awareness and citizen engagement [32].

Dominance metrics provided additional insight into community structure. For birds, Cumulative Dominance remained stable, oscillating between 46% (2022) and 53% (2023), while Simpson’s Index of Dominance increased modestly from 0.032 at baseline to 0.051 in 2023, easing slightly in 2024 (0.048). Collectively, these values indicate that roughly half of all individuals were consistently concentrated in a small suite of disturbance-tolerant species, mirroring the gradual consolidation already suggested by the increases in average population size. Insects showed consistently low dominance during the monitored period. Cumulative Dominance never exceeded 30%, while Simpson’s Index of Dominance remained tightly clustered around 0.03. These figures indicate that no single species or small set of species monopolized the assemblage once systematic sampling was in place; instead, individuals were dispersed across a broad taxonomic spectrum, yielding a markedly even community that contrasts with the more hierarchical structure observed in birds.

Comparisons across years revealed that the most frequently recorded species in 2022 was Serinus serinus Linnaeus, 1758, (n = 53), while in both 2023 and 2024 it was Erithacus rubecula Linnaeus, 1758 (n = 94 each year). The most notable change was observed in point 5, where the dominant species shifted from Hirundo rustica Linnaeus, 1758 in 2022 (n = 9) to Passer domesticus Linnaeus, 1758 in 2024 (n = 16).

The increase in bird abundance despite a decline in species richness is a noteworthy outcome. This pattern, where a few generalist species dominate, is often seen in urbanized environments due to habitat homogenization [18]. For instance, E. rubecula and P. domesticus both showed significant increases in abundance, reflecting their adaptability to altered urban habitats [33]. Their success can be linked to flexible foraging strategies and broad nesting preferences that allow them to exploit a wide range of urban resources [34].

As shown in Figure 3, the average bird population size increased across most observation points from 2022 to 2024, except for point 5, which experienced a decline in 2023 before rebounding in 2024. Figure 4 further illustrates a consistent yearly increase in the number of individuals per point from 2022 to 2024. The baseline was excluded from this comparison.

Vegetation characterization prior to bird and insect surveys showed the following patterns: (a) the UGA-A (Penafiel City Park) featured ornamental trees and shrubs, arranged in a garden layout with mown grass; (b) the UGA-B had transitional vegetation between the City Park and valley, composed of oak woodland, small shrubs, and tall herbs; (c) the UGA-C was dominated by shrubs and tall conifers, with alien species such as Acacia spp., meadows, and traces of peri-urban agriculture; and (d) the UGA-D was predominantly a riparian zone, with adjacent meadows, a more naturalized habitat, some alien species, and several small waterfalls.

The vegetation type can influence bird and pollinator presence, depending on plant structure (trees, shrubs, grasses) and bird behaviours such as nesting and foraging. Insectivorous species may not depend on trees for nesting but can be influenced by shrub presence, as many insects and pollinators rely on such vegetation [35]. In Salamanca, Peris and Montelongo [36] reported a positive correlation between Carduelis chloris Linnaeus, 1758 and S. serinus abundance and the presence of trees and shrubs in urban parks, similar to point 2 in this study, where S. serinus was the most recorded species in 2023 (n = 26). The presence of the Cavalum River may also help explain the relatively stable richness values, as riparian corridors are commonly used by species at the landscape scale [37,38]. Several forest-associated species, like Parus major Linnaeus, 1758, Cyanistes caeruleus (Linnaeus, 1758), Fringilla coelebs Linnaeus, 1758 and, Turdus merula (Linnaeus, 1758), were found throughout the valley, supporting findings by Jokimäki and Suhonen [39] that these species have partially adapted to urban environments. Trophic interactions play a crucial role in these dynamics: birds, as secondary and tertiary consumers, rely on insects as a food source. Hence, increases in insect abundance could benefit bird populations, through the concurrent decline in insect species richness may signal a loss of specialists, potentially affecting long-term ecological resilience and function [40]. In Shanghai’s urban core, Wang, et al. [41] demonstrated that corridors exceeding 2 ha with adjacent low-traffic greenways supported 25% higher bird richness than smaller, isolated patches. These insights underscore why our combined planting and invasive removal interventions must be evaluated both at the stand scale and across the landscape level.

According to Paquet [42], who compiled a list of 37 bird species commonly found in 16 European cities, the Cavalum Valley hosts 23 of these. Urban-adapted species such as Sturnus vulgaris Linnaeus, 1758 and P. domesticus [43] were consistently present each year. Furthermore, of the top 10 indicator species of high environmental quality in European cities identified by Morelli, et al. [44], 7 were frequently observed in the valley: C. caeruleus (Linnaeus, 1758), F. coelebs, P. major, Pica pica Linnaeus, 1758, S. vulgaris, Sylvia atricapilla Linnaeus, 1758, and, T. merula. These species benefit from the diverse food resources and nesting opportunities provided by urban green spaces. The urban success of S. vulgaris aligns with findings by Southon, et al. [45], who documented similar patterns across European cities.

Mason [46] reported that while bird richness was similar between urban and rural areas, bird abundance was significantly lower in nearby farmland. In the present study, sampling point 1 (more urban) showed higher bird richness and abundance in both 2023 and 2024 compared to sampling point 5, which features more naturalized conditions. This may also reflect anthropogenic disturbances. These results are in line with studies by Pennington, et al. [47], who identified vegetation complexity as a key driver of bird diversity in urban landscapes. The UGA-B, a transition area between City Park and oak forest, supported high richness of both birds and plants, highlighting the importance of heterogeneous habitats in sustaining biodiversity [48]. Similarly, in Nanjing, China, bird abundance was greater in larger parks than in smaller ones [49], which is consistent with observations from Penafiel City Park (UGA-A).

3.2. Response off Functional Groups and Indicator Species

Analysis of bird functional groups (

Table 5. Species richness by functional group for birds.

Category	Group	Richness (n=)
Foraging	Ground	21
	Shrubs	10
	Trees	10
	Air	13
	All	9
Nesting	Burrow	9
	Tree Branches	17
	Shrubs/Bushes	14
	Hollow	11
	Artificial	5
	Other	7
Diet	Carnivores	4
	Granivore	15
	Insectivore	34
	Omnivore	10

) revealed that ground foraging was the most common foraging strategy. Most bird species were found to nest in tree branches and bushes (n = 17 and n = 14, respectively). Dietary analysis highlighted strong dominance of insectivores, with 34 species representing 53.97% of all recorded bird species. Among carnivorous birds, 75% nested in tree branches. Notably, of the 21 ground-foraging species 47.61% were insectivores.

The presence of birds of prey such as Accipiter nisus Linnaeus, 1758, and Buteo buteo Linnaeus, 1758, is not unexpected given that these raptors are increasingly using urban areas to hunt small birds and mammals that thrive in cities. These prey species, including seagulls, doves, and rodents, are often regarded as urban pests [50]. Preserving and enhancing urban habitats suitable for birds of prey may contribute to natural pest control in urban ecosystems.

Bayulken, et al. [51], in a study conducted in Melbourne, noted that urban greening and landscaping around buildings often involve a mix of non-native ornamental plant species chosen more for esthetic or maintenance purposes than for ecological function. Nonetheless, such areas still provide food, shelter, and nesting opportunities for birds and other urban-adapted fauna. The Cavalum Valley, therefore, may offer a significant ecological value to the city, especially considering the presence of species with known human associations, such as Periparus ater Linnaeus, 1758, which often visits feeders, and Cyanistes caeruleus Linnaeus, 1758, which commonly nests in artificial nest boxes [35,52]. Nesting preferences further elucidate the relationship between birds and vegetation structure. The predominance of species nesting in tree branches underlines the importance of maintaining mature trees and varied arboreal structures in urban green spaces. These findings are consistent with Marzluff, Bowman and Donnelly [33], who emphasized that urban areas with abundant tree cover tend to support higher bird diversity. In contrast, ground- and shrub-nesting birds were less represented, likely due to higher predation risk and human disturbance in urban contexts [53].

The findings of this study align with research from other cities implementing NbS [17,22,51,54,55]. For example, urban greening initiatives in London have led to measurable increases in bird abundance and pollinator diversity, highlighting the potential of well-designed green infrastructure to counteract biodiversity loss [48]. However, the success of such interventions depends on local factors, such as climate, landscape context, and urban form, which necessitate site-specific approaches to green space planning.

Regarding insect species, Lepidoptera was the most frequently recorded order (Figure 5), with 21 species in 2022, 22 in 2023, and 27 in 2024. Across all three years, the most abundant species was Apis mellifera Linnaeus, 1758, with population counts of 46 in 2022, 72 in 2023, and 91 in 2024. Pollinators accounted for 37.95% of insect records in 2022, increasing to 48.4% in 2023 and 51.3% in 2024, indicating a positive trend in pollinator representation. Insect species with conservation status under Annex II of the EU Habitats Directive (Cerambys cerdo Linnaeus, 1758, Coenagrion merculiale Charpentier, 1840, Euplagia quadripunctaria Poda, 1761, and Lucanus cervus Linnaeus, 1758) also showed an increase in total abundance, reaching a combined total of 17 individuals in 2024. However, the average population size per species remained below expected levels, suggesting limited recovery or the need for targeted conservation actions.

3.3. Effectiveness of Management and Nature-Based Solutions: Temporal Trends and Ecological Patterns

Across the study area, although insect species richness declined slightly in 2023 after the NbS implementation, the average insect population size increased from 9.25 in 2022 to 12.20 in 2024 (Table 4). This was accompanied by an overall increase in insect abundance, mainly driven by pollinators species. This pattern suggests that certain insect taxa have benefited from habitat changes, likely due to alterations in vegetation structure and floral resource availability. For example, the continued dominance of A. mellifera indicates that the UGAs provide suitable conditions for pollinators, such as abundant and diverse floral resources essential for foraging [56]. The increase in pollinator abundance may generate cascading ecological benefits, including enhanced plant reproductive success and improved biodiversity, thereby strengthening the ecological function of urban green spaces [57]. In urban fringes, pollinators play a vital role in supporting vegetation communities through pollination and serve as key prey for higher trophic levels, contributing to the overall urban ecosystem health and resilience [58].

The Shannon–Wiener Diversity Index for insects showed a minor reduction, from 3.93 in 2022 to 3.85 in 2024, suggesting slight shifts in species composition and evenness rather than a substantial biodiversity loss. While urbanization typically threatens biodiversity, recent studies [59,60] confirm that pollinator groups, especially wild bees, can thrive in urban environments. Local factors (e.g., native floral availability) and landscape-level characteristics (e.g., proportion of impervious surfaces) strongly influence pollinator community structure. Conversely, urban intensification, chemical use, climate change, and competition with high-density honeybee colonies exert detrimental effects on urban pollinator populations [60]. Moreover, considering the presence of insectivorous birds, predation pressure may also impact pollinator dynamics [61]. While Udy, et al. [62] reported that bumblebees were the dominant floral visitors in urban areas, A. mellifera was clearly the most abundant pollinator in the Cavalum Valley. This may partly reflect the influence of nearby agricultural areas (particularly near the UGA-C), which likely contribute to honeybee population surges.

In terms of vegetation, 191 plant species were recorded in the Cavalum Valley, distributed by 72 botanical families. Fabaceae was the most abundant family (n = 12), although 25% of its species were non-native (Acacia spp.). Only 5.23% of all recorded species held the RELAPE (Rare, Endemic, Localized, Threatened, or Endangered) status, including Ilex aquifolium L., Quercus suber L., Omphalodes nitida Hoffmanns. & Link, and Echium rosulatum Lag.). These species were most frequent in Spot 4, located within the UGA-C (Figure 6).

Vegetation structure also changed temporally. In the UGAs B, C, and D, herbaceous layers remained dominant from 2023 to 2024. Although the shrub layer was initially scarce, it gradually expanded over the study period. The tree cover remained stable overall, even as new trees were planted. Notably, species such as Urtica dioica L. in Spot 5 declined significantly, from a coverage class of [51%; 75%] in 2022 to [6%; 25%] in 2024. In Spot 1, invasive species like Acacia spp. and T. fluminensis Vell. were successfully reduced, making room for native and RELAPE species, such as Linaria triornithophora (L.) Willd. and Foeniculum vulgare Mill. In the UGA-A, the number of plant species remained at 70 in 2022 and increased to 71 in 2023, following the plantation of Crataegus monogyna Jacq. Nearby the UGA-B. Although the ground layer remained the most frequent vegetation stratum, the tree layer was prominent.

As vegetation height and density increased, it facilitated the formation of multi-layered habitats, which in turn supported greater diversity of bird species. This aligns with earlier findings [12,19,38] confirming that more complex vegetation structures offer increased foraging and nesting opportunities for birds. Additionally, insects, like solitary bees and certain wood-boring beetles help create secondary habitats for other species. For example, wood-boring beetles create holes in trees that can be used by birds for nesting [63]. The presence of a riverine ecosystem (Figure 7) and adjacent forest patches also contributed to insect diversity, especially pollinators, particularly in the UGA-B, where bird abundance was relatively lower. The reduced bird predation pressure may have facilitated insect population growth in this zone.

Overall, these findings highlight the interconnectedness of vegetation structure, faunal diversity, and ecological processes. They reinforce the importance of maintaining structurally heterogeneous and botanically diverse green spaces as part of urban conservation strategies. The analysis of bird functional groups and community composition highlights the importance of diverse and well-managed green infrastructure in supporting ecosystem resilience. Urban green areas like those in the Cavalum Valley can effectively sustain diverse avian guilds by supplying essential resources, from food to nesting substrates. The observed trends are consistent with studies from other cities implementing NbS [13,35,36,64], further supporting the value of functional diversity in promoting ecological stability [44].

Europeans cities are already implementing NbS. Strategies include regreening or creating biodiverse urban spaces, which can help combat climate change, manage water resources, and provide access to nature for urban residents [65]. Spatial modelling frameworks can guide urban planners in prioritizing green infrastructure development to support both pollinator habitats and human health, particularly in areas of high deprivation [66]. While NbS are increasingly adopted in European cities, there is still untapped potential for implementation at the building and district levels [67], where our case study can provide some insight. These approaches contribute to urban resilience, biodiversity conservation, and the creation of more sustainable and attractive urban environments.

Figure 8 illustrates variation in vegetation layers, bird foraging type, and insect ecological roles across the UGAs between 2022 and 2024. The UGA-C exhibited high bird and vegetation richness, corresponding to a more complex vegetative structure. In contrast, insect richness often showed an inverse trend, particularly evident in the UGA-B, where lower vegetation and bird richness coincided with higher insect diversity. This is likely due to the presence of the oak forest and open meadows, which provide optimal conditions for pollinators.

4. Implications for Restoration and Management

The documented positive responses of birds and insect communities to vegetation management and NbS interventions underscore the value of integrating biodiversity considerations into ecosystem restoration. The increase in indicator species and functional groups demonstrates that targeted vegetation structure improvements and native species planting can rapidly enhance habitat quality and ecosystem services (Figure 9).

Practically, the findings support the prioritization of mixed native shrub and tree species to create heterogeneous structural mosaics, which foster diverse faunal assemblages. The promotion of flowering plants benefits pollinator populations, which are critical for both natural ecosystems and adjacent agricultural areas.

Furthermore, monitoring programmes should incorporate indicator taxa to provide cost-effective and sensitive assessments of restoration progress. The observed functional shifts suggest that NbS can contribute to trophic complexity and ecological resilience, reinforcing their role in sustainable landscape management.

Managers should also consider the temporal scale of restoration, maintaining long-term commitments to vegetation maintenance and monitoring to ensure sustained ecological benefits. Adaptive management, informed by ongoing ecological data, will be essential to address emerging challenges such as invasive species and climate change impacts.

A key outcome of our work is the demonstrated interdependence among the three biological groups—birds, pollinators, and plants—which underpins the ecological dynamics within the study area. The increased structural complexity and floral resource availability following the NbS implementation provided essential foraging and nesting habitats, supporting both insect pollinators and insectivorous birds. Pollinators, particularly bees and Lepidoptera, were closely associated with the herbaceous and shrubby layers, facilitating plant reproduction and enhancing vegetation diversity, which in turn supported a broader array of bird functional groups. Insectivorous birds contributed to regulating herbivore populations, creating feedback loops that influenced both floral community structure and pollinator dynamics. These trophic and mutualistic interactions underscore the importance of integrated multi-taxon assessments, since shifts in one group can cascade through others and ultimately affect ecosystem function and resilience.

Functional-group analyses further emphasized that vegetation structure is critical in supporting bird foraging and nesting behaviours: stratified vegetation layers correlated with higher species richness and trophic complexity. Although some trade-offs were observed, such as temporary reductions in particular taxa or shifts in insect community composition, the overall trend indicated synergistic biodiversity gains across taxa. This study also illustrates the utility of rigorous, field-based monitoring to capture dynamic ecological responses to NbS. The combined use of birds, pollinators, and flora as complementary bioindicators provided a nuanced understanding of recovery processes, and the observed trajectories of recovery affirm that strategically designed, ecologically grounded NbS can foster multifunctional, biodiverse, and resilient urban green spaces.

Moreover, our findings contribute directly to several United Nations Sustainable Development Goals: SDG 11 (Sustainable Cities and Communities) through the enhancement of urban green infrastructure; SDG 13 (Climate Action) via ecosystem-based mitigation and adaptation; and SDG 15 (Life on Land) by promoting terrestrial biodiversity conservation and ecosystem restoration. By supporting these goals, the study reinforces the critical role of NbS in sustainable urban development. Ongoing long-term monitoring and adaptive management are essential to sustaining biodiversity gains and addressing emerging pressures such as urban expansion and climate variability.

5. Conclusions

This study aimed to evaluate the effectiveness of NbS in enhancing biodiversity within urban green spaces by monitoring ecological changes across riverine and forested habitats in the Cavalum Valley, Penafiel. Through systematic multi-taxon field surveys conducted over three years, our findings demonstrate that integrated NbS interventions can substantially contribute to biodiversity restoration in urban landscapes. Results showed that NbS led to significant increases in the abundance and average population size of both birds and pollinators despite initial declines in species richness following disturbances such as invasive species removal. By the third year, biodiversity metrics had stabilized or improved beyond baseline values, indicating a resilient ecological response. Furthermore, the data revealed differential impacts depending on the NbS type: afforestation and riparian restoration notably benefited avifauna, while enhancements in floral resources supported higher pollinator abundance, particularly Apis mellifera.

Together, these outcomes confirm that well-designed, multi-layered NbS can rapidly enhance multi-taxon biodiversity and habitat complexity in Mediterranean-climate cities, offering a robust pathway for simultaneous ecological and social benefits.

Future research should therefore prioritize identifying ecological thresholds of resilience, enhancing community engagement in biodiversity stewardship, and testing the scalability of NbS interventions across diverse urban contexts.