Introduction
Lichens have long served as indicators of air pollution, as detailed in a review by Hawksworth (1970). Research throughout the 1970’s primarily focused on toxic effects of SO2 on lichen bionts (Rao and LeBlanc 1966; Nash III 1973) and on the threat SO2 posed to lichen communities in urban and industrial areas (LeBlanc and Sloover 1970; Hawksworth et al. 1973). In subsequent years, growing concerns emerged that H2SO4, which forms from SO2 and is deposited as acid rain, could indirectly harm lichens by acidifying their substratum (Farmer et al. 1991; Gauslaa 1995). The deposition of sulfur (S) peaked in northern Europe in the late 1960’s, followed by a rapid and substantial reduction in acid rain (e.g., Fowler et al. 2007; Grennfelt et al. 2020). However, the deposition of nitrogen (N) remained high or showed only weak declines (Aas et al. 2017). The surge in diesel vehicles, a recent contributor to nitrogen oxides (NOx) in Norway, has impeded effective control measures for N-pollution (Tønnesen 2010). Currently, airborne NOx and NHx continue to pose significant environmental challenges (Rockström et al. 2009) with the potential to affect lichens (Hauck 2010) and disrupt lichen vegetation (Pinho et al. 2012; Carter et al. 2017; Esseen et al. 2022). Excess N may favor nitrophytic lichens (Munzi et al. 2019) at the expense of other species (van Dobben and ter Braak 1999; Frati et al. 2007). Interestingly, nitrophytes thrive in conditions of elevated pH (e.g., De Bakker 1989; van Herk 1999; 2001), suggesting that the functional triggers for these lichens are not fully understood.
The detrimental impacts of air pollution on lichens were first observed in urban environments. Sernander (1926) applied a three zone system based on epiphytic lichens. He referred to the inner lichen-deficient city as a “lichen desert”, surrounded by the “struggle zone” that hosted an impoverished lichen flora with compromised viability, while the “normal zone” in the city’s outskirts supported healthy epiphytic lichen communities. Oslo was among the first cities where lichens were mapped and where these zones were illustrated (Haugsjå 1930). Throughout the period from 1973 to 2019, which encompassed successive pollution regimes, lichen ecology students from the Norwegian University of Life Sciences recorded epiphytic lichens under the author’s supervision. Results from the 1970’s were summarized in an unpublished MSc-thesis (Øyseth and Aarvik 1980) shortly after the peak in S-deposition and acid rain. Data from subsequent years, characterized by rapidly declining S-deposition and high, but slightly decreasing N-deposition (Aas et al. 2017) have yet to be compiled. In total, lichen species richness in Oslo has been documented over nearly a century during which the composition of airborne pollutants successively changed. This provides a unique opportunity not only to meet a need for including lichens in long-term ecological studies (Diekmann et al. 2023), but also to examine changes in epiphytic lichen richness over the past 90 years along an urban-to-rural gradient. By comparing the changes in richness of the nitrophytes and acidophytes in the context of strongly declining acid rain yet persistently high N-depositions, we may enhance our understanding of the factors that shape these two respective functional groups of lichens.
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The first objective of this study is to quantify changes of epiphytic macrolichen species richness in Oslo during decreased pollution from 1973 to 2019, which is the focal period for the fieldwork. The aim is to establish a time scale for lichen recovery, thereby evaluating the hypothesis that lichens not only can monitor sequential escalations but also declines of air pollution levels. By making these comparisons, we test the hypothesis that previous lichen vegetation can recolonize urban areas after the pollution-induced lichen dieback in the 1970’s.
The second objective is to compare responses of acidophytes (lichens in the Pseudevernion community associated with oligotrophic bark) and nitrophytes (species in the Xanthorion associated with relatively high bark pH and/or excess N; Du Rietz 1945; Barkman 1958; van Herk 2001) across temporal and spatial gradients. The underlying hypothesis posits that acidification will lead to a decrease in nitrophytes and an increase in acidophytes, with the reverse occurring during periods of declining acid rain. Ultimately, the findings could provide valuable insights for predicting future trends in epiphytic lichen vegetation in populated landscapes where high N-deposition is likely.
The final objective, which will be outlined in the discussion, is to compare the epiphytic lichen vegetation from 1930 (Haugsjå 1930) during the early rise in S-deposition with the two subsequently reported periods: (1) shortly after the peak of S-deposition in the 1970’s and (2) four decades later in the 2010’s, a period characterized by low S-deposition and reduced acid rain. Given that lichens serve as spatially transferable pollution bioindicators (Delves et al. 2023), the documented trends should be applicable in other places as well.
Materials and methods
Study area and air pollution trends
The study was conducted in the boreonemoral zone forming a transition between the boreal and nemoral zones (Moen 1999) in southeast Norway. A 5–8 km broad section extending southeastwards from Oslo city (59°55′1.22″N, 10°43′44.77″E) to rural environments (59°39′57.86″N, 10°46′0.32″E) was examined. In the 1970’s, the section ended 10 km from the city center, which then had rural settings and normal lichen flora. As the city and concurring urban pollution expanded, the gradient was extended to 27 km from the city center in 1998 to encompass intact rural settings in following years. The gradient comprised 104 sites (parks, graveyards, and other semi-open areas) situated below 200 m a.s.l.
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Field work started in 1973, shortly after the peak in S-deposition, which was associated with large-scale damage to the lichen flora (Hawksworth and Rose 1970). This peak in S-deposition (e.g., Lee 1998) was swiftly followed by a rapid decline (Vestreng et al. 2007). The yearly mean SO2-concentration across Oslo declined from ≥ 100 µg SO2 m− 3 (1963–1970; the highest site mean ~ 250 µg SO2 m− 3) to ≤ 2 µg SO2 m− 3. Moreover, no sites had concentrations exceeding 4 µg SO2 m− 3 after 1994 (Gram 2005). This period of high S-deposition coincided with high N-deposition (Aas et al. 2017). However, N-deposition stayed high with only minor declines (Wright et al. 2001). The yearly mean NOx concentration across sites in Oslo was 49–55 µg m− 3 in 1960–1992, which dropped to 40 µg m− 3 in 1998 (Gram 2005). NHx is a minor component in Oslo (Grøntoft 2021), likely due to the city’s location between the sea and forested mountains, with no nearby agriculture. Diesel vehicles have recently emerged as a significant contributor to N. This, along with an increase in traffic volume, has resulted in the stagnation of the decline in urban NOx (Tønnesen 2010).
Field work
The field work was done annually by groups of 3–4 student as part of a term assignment in lichen ecology at the Norwegian University of Life Sciences. Each group was given 3–4 localities, with distances progressively increasing from the city center, which was defined by the location of the parliament building. Within each year, all groups studied unique sites, although many of these sites had previously been investigated by other groups in past years. In total, 411 investigations were done at 104 sites over a span of 47 years.
Prior to the field work, the students underwent a comprehensive training of 6 h spread over three weeks. This training focused on the identification of epiphytic lichens with the aid of microscopes and chemicals for color tests. Following the guided training, each group visited their sites in September and were instructed to select the three most lichen-rich tree trunks in each site of (1) broadleaved deciduous trees (mainly Fraxinus, Populus, Ulmus, Acer, Tilia) and (2) coniferous trees and/or Betula species. For each selected trunk, the students listed all macrolichen species seen to a height of 2 m. They collected one specimen of all species by cutting the underlying dead outer bark layer from each of the two categories of trees. After the field work, the students identified their specimens in the lab, guided by their teacher.
No information of earlier results was given to the students prior to field work. This means that for sites that had been visited a previous year, a new group did not have a list of expected lichens and most likely did not examine the same tree trunks chosen by earlier groups. Therefore, each of the 411 studies conducted could be regarded as unique and independent.
Data retrieved from literature
Distribution of epiphytic lichen species was mapped in Oslo in 1930, based on species lists from 126 sites (Haugsjå 1930). This data serves as a benchmark for macrolichen richness before the main onset of acidification. The published species lists show that a variety of tree species were investigated at each site. For this study, species lists from 107 sites were used. The sites that were excluded sites were either situated > 200 m a.s.l. or had location names that prevented accurate calculation of distances from the city center.
Taxonomic challenges: The lichen Physconia grisea, a rarity in Norway, was the only sorediate Physconia taxon recorded in 1930. Here, it is referred to as P. enteroxantha, which is the common sorediate Physconia species, and likely encompassed P. perisidiosa in the 1930 records. The only Usnea species recorded in that year was U. barbata, which, with the benefit of current knowledge, was primarily U. dasopoga. The name Bryoria jubata used in 1930 included B. fuscescens and B. capillaris. Lastly, the following species, mainly rare in 1973–2019, were not recognized or recorded by Haugsjå (1930): Cetraria sepincola, Hypogymnia farinacea, Imshaugia aleurites, Parmeliopsis hyperopta, Umbilicaria hirsuta, Ramalina fastigiata, Phaeophyscia endophoenicea, P. nigricans, P. sciastra, Physcia subalbinea, and Pleurosticta acetabulum. Due to uncertainties whether these species were recognized, the species richness in 1930 was probably underestimated.
Lichen communities
The sampled tree categories, namely broadleaved trees and conifers/Betula spp., have bark that is classified as base cation-rich and acidic, respectively (Du Rietz 1945). These trees typically host distinct epiphytic chlorolichen communities (Barkman 1958): the Xanthorion (mostly foliose members of the Teloschistales and the Physciaceae), and the Pseudevernion (most foliose members of the Parmeliaceae), respectively. This division aligns with the distinction between nitro- and acidophytes (e.g., van Dobben et al. 2001). However, these communities are not strictly confined to their typical tree hosts. For instance, conifers can host the Xanthorion in the presence of cation-rich dust and/or N-compounds (e.g., Gurholt 1968). The different chemical preference of Xanthorion and Pseudevernion was supported by a study in spruce canopies in British Columbia (Gauslaa et al. 2021a) where Xanthorion species were associated with high bark pH and high concentration of base cations in the canopy throughfall, whereas Parmeliaceae species grew on acidic bark exposed to Mn-rich and base cation-poor canopy throughfall. The brown Parmeliaceae genera Melanelixia and Melanohalea, which associate with an intermediate elemental composition (van Dobben et al. 2001), were included in the Pseudevernion due to their frequent occurrence on conifers and Betula. The Usneion community was dominated by fruticose lichens such as Ramalina on less acidic bark and Bryoria on trees with low pH and acidic canopy throughfall (Gauslaa et al. 2021a). Finally, a few cyanolichens, occasionally found only on broadleaved trees, characterized the Lobarion community (Gauslaa 1985).
Verification and selection of species data
Only macrolichens were recorded because many crustose lichens could not be accurately identified. Sorediate Polycauliona species were referred to as P. candelaria, and isidiate Parmelia species as P. saxatilis. The Candelaria found was referred to as Candelaria pacifica (Westberg and Arup 2010) but may include specimens of C. concolor. In 1973–1979, all small Usnea specimens that could not be identified morphologically, were subjected to thin-layer chromatography (TLC) by Øyseth and Aarvik (1980) with methods outlined by (Menlove 1974). The resulting chemical data revealed that all small Usnea specimens in the city were U. dasopoga, while U. hirta and U. subfloridana only occurred in rural areas where specimens were well developed (Øyseth and Aarvik 1980). Consequently, all Usnea collections that could not be morphologically identified were labelled as U. dasopoga. Bryoria specimens that tested negative for KOH were identified as B. fuscescens, which includes B. vrangiana, B. pseudofuscescens, and B. implexa according to TLC data (Øyseth and Aarvik 1980). Conversely, specimens with a positive KOH-test were identified as B. capillaris. Cladonia species were often present but were not identified or recorded.
The author examined every collected sample including the adjacent bark and either validated or corrected the students’ species identification. Overlooked species in collected samples were added to the final list. As a result, the dataset solely comprised checked and verified taxa. To ensure robust data, only the presence of species on sampled trees was used. Species richness refers to the total number of macrolichens found on the six trees studied at each site. Each species list thus represented an independent observation of lichens from newly selected trunks at each site and year, conducted by a new group of students with a standardized training. To eliminate potential artifacts due to teacher-dependent instruction and species knowledge, collections from the years when the author was on research leave were excluded. Data from the years 1980–1997 had not been secured and were therefore omitted. Thanks to careful checking of all collections, the probability of falsely reporting too many species was minimal.
Statistical analyses
Statistical analyses were run in Minitab® 21.4.1. Linear regression analyses were performed to assess the relationship between species richness and (1) the log-transformed distance from the city center, and (2) time (year) within the most frequently recorded central urban park urban. A two-way ANOVA was conducted to examine the relationship between species richness and (1) two lichen communities (Xanthorion and Pseudevernion), and (2) seven time periods (1973–1975, 1975–1979, 1998–2000, 2001–2005, 2006–2010, 2011–2015, 2016–2019). These periods were selected to encompass a similar number of years, while considering that some years lacked data and to differentiate the period before and after the year 2000 when a weather-induced dieback occurred (Gauslaa 2023). The datasets met the requirements for the respective models. Trendlines, along with standard error bands, for macrolichen richness versus distance in 1930, 1973–1979, and 2013–2019 were plotted using Jamovi (2022).
Results
Macrolichen richness across urban to rural regions 1973–2019
55 macrolichen species were found across years and sites (Table 1). The chlorolichens were divided into three categories: 23 foliose species were classified as Xanthorion species (nitrophytes), 20 as Pseudevernion species (acidophytes), and the 9 fruticose species as Usneion species (Table 1). Three cyanolichens (Lobarion) were occasionally found. The most frequent species (≥ 50%) included acidophytes such as Parmelia sulcata, Hypogymnia physodes, Pseudevernia furfuracea, and nitrophytes like Physcia tenella, Xanthoria parietina, Phaeophyscia orbicularis (Table 1).
Table 1
Macrolichens found in Oslo and surroundings in 411 studies (6 trunks in each study) from 1973 to 2019, ranked after decreasing occurrences (last column). The three first numeral columns show the percent frequency of lichens in sites within 10 km from the city in 1930 (data from Haugsjå 1930; n = 107), the first (1973–1979; n = 46), and the last 7-year period (2013–2019; n = 111) of this study, respectively
Species | Community | Percent occurrence | Total n (of 410) | ||
|---|---|---|---|---|---|
19301 | 1973-79 | 2013-19 | 1973–2019 | ||
Parmelia sulcata | Pseudevernion | 82.2 | 42.6 | 89.2 | 336 |
Hypogymnia physodes | Pseudevernion | 68.2 | 66.0 | 66.7 | 302 |
Physcia tenella | Xanthorion | 42.1 | 14.9 | 82.9 | 286 |
Xanthoria parietina | Xanthorion | 69.2 | 19.1 | 73.9 | 264 |
Phaeophyscia orbicularis | Xanthorion | 59.8 | 10.6 | 78.4 | 247 |
Pseudevernia furfuracea | Pseudevernion | 12.2 | 23.4 | 34.2 | 202 |
Polycauliona polycarpa | Xanthorion | 49.5 | 8.5 | 45.9 | 192 |
Hypogymnia tubulosa | Pseudevernion | 0.9 | 4.3 | 41.4 | 166 |
Platismatia glauca | Pseudevernion | 18.7 | 21.3 | 27.9 | 165 |
Melanelixia glabratula | Pseudevernion | 30.0 | 48.6 | 159 | |
Melanohalea exasperatula | Pseudevernion | 44.9 | 12.8 | 49.5 | 147 |
Tuckermannopsis chlorophylla | Pseudevernion | 25.2 | 36.2 | 19.3 | 142 |
Evernia prunastri | Usneion | 29.0 | 8.5 | 18.0 | 139 |
Parmeliopsis ambigua | Pseudevernion | 17.8 | 38.3 | 33.3 | 138 |
Physconia enteroxantha | Xanthorion | 43.0 | 8.5 | 36.0 | 130 |
Vulpicida pinastri | Pseudevernion | 20.6 | 40.4 | 23.4 | 121 |
Polycauliona candelaria | Xanthorion | 44.9 | 2.1 | 36.9 | 118 |
Bryoria fuscescens | Usneion | 14.9 | 29.8 | 18.0 | 117 |
Physconia distorta | Xanthorion | 59.8 | 8.5 | 16.2 | 109 |
Physcia stellaris | Xanthorion | 55.1 | 12.8 | 36.9 | 101 |
Usnea dasopoga | Usneion | 14.0 | 10.6 | 8.1 | 94 |
Physcia dubia | Xanthorion | 79.4 | 23.4 | 29.7 | 94 |
Parmelia saxatilis | Pseudevernion | 3.7 | 6.4 | 16.2 | 86 |
Physcia adscendens | Xanthorion | 31.8 | 4.3 | 36.0 | 82 |
Melanelixia subaurifera | Pseudevernion | 9.3 | 19.8 | 73 | |
Ramalina farinacea | Usneion | 1.9 | 2.1 | 4.5 | 64 |
Parmelina tiliacea | Xanthorion | 26.2 | 4.3 | 3.6 | 58 |
Anaptychia ciliaris | Xanthorion | 20.6 | 2.1 | 0.0 | 56 |
Phaeophyscia nigricans | Xanthorion | 0.0 | 27.0 | 45 | |
Physcia aipolia | Xanthorion | 32.7 | 4.3 | 6.3 | 33 |
Candelaria pacifica | Xanthorion | 84.1 | 0.0 | 15.3 | 31 |
Bryoria capillaris | Usneion | 4.3 | 3.6 | 27 | |
Hypogymnia farinacea | Pseudevernion | 0.0 | 4.5 | 22 | |
Peltigera praetextata | Lobarion | 0.0 | 2.7 | 22 | |
Phaeophyscia sciastra | Xanthorion | 0.0 | 8.1 | 19 | |
Usnea hirta | Usneion | 0.0 | 0.0 | 13 | |
Physconia perisidiosa | Xanthorion | 0.0 | 2.7 | 13 | |
Ramalina fraxinea | Usneion | 5.6 | 0.0 | 0.0 | 12 |
Melanohalea exasperata | Pseudevernion | 11.2 | 0.0 | 4.5 | 11 |
Cetraria sepincola | Pseudevernion | 8.5 | 0.9 | 10 | |
Parmeliopsis hyperopta | Pseudevernion | 8.5 | 0.9 | 6 | |
Physcia subalbinea | Xanthorion | 0.0 | 3.6 | 6 | |
Physcia caesia | Xanthorion | 15.0 | 0.0 | 3.6 | 6 |
Pleurosticta acetabulum | Xanthorion | 2.1 | 0.0 | 6 | |
Melanelixia subargentifera | Pseudevernion | 5.6 | 0.0 | 2.7 | 5 |
Imshaugia aleurites | Pseudevernion | 0.0 | 0.0 | 5 | |
The comprehensive dataset on species richness observed in this study, depicted in 3-D graphs by functional lichen groups with year and distance from the city center as axes, is given in Fig. 1a-d. In the 1970’s, macrolichens were not found in sites within 1 km of the city center (Fig. 1a), while fruticose lichens (Usneion) were absent within a 4 km radius (Fig. 1b). Acidophytes (Fig. 1c) were most common (e.g., Parmeliopsis ambigua, Platismatia glauca, Pseudevernia furfuracea, Tuckermannopsis chlorophylla, Vulpicida pinastri; Table 1). Thereafter, a variety of species, including fruticose lichens, re-established (Table 1). After 1998, macrolichens were observed at all sites in the city (Fig. 1a). For instance, the number of macrolichens at the central urban park around Oslo Cathedral rose from 0 in 1977 to 7 species in 2019 (number of lichens = -358.5 + 0.181 x year; R2adj = 0.605; P < 0.001; n = 18 years; linear regression).
Fig. 1
The richness of a: total macrolichen species, b: Usneion species, and c: Pseudevernion species, and b: Xanthorion species in 411 studies (6 trunks in each study) across time (year) and distance from the city center. The species included in the three epiphytic lichen communities are specified in Table 1
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Following a period devoid of reported studies (1980-97), the years 1998 to 2000 represented a temporal peak in macrolichen richness with notably higher number of species compared to the 1970’s (Fig. 1a). However, this trend abruptly reversed between 2001 and 2005, before a second surge in macrolichen richness commenced, continuing until the conclusion of the study in 2019. This fluctuation in species richness impacted all three communities similarly, although the temporal decline in species was most pronounced outside the city (Fig. 1b-d).
Urban-suburban trends in acidophyte and nitrophyte richness
The Xanthorion species (Fig. 1d) exhibited a stronger response to time, as compared to Pseudevernion species, which were more responsive to the distance from the city center (Fig. 1c). In the following, the species richness of these two groups was scrutinized within the urban-suburban gradient, specifically focusing on sites closer than 10 km from the city center. A comparison of species in the first (1973–1979) and last 7-y period (2013–2019) (Table 1) revealed that 16 out of the 20 most common lichens, present in > 100 studies, experienced an increase in frequency. The frequency of the three most common nitrophytes (Physcia tenella, Xanthoria parietina, Phaeophyscia orbicularis) increased from ≤ 19% in initial period to as much as ≥ 74% in the concluding period (Table 1). In contrast, no acidophytes achieved such high increase rates. Only four out of the 20 most common species decreased in frequency between these two periods. All four (Tuckermannopsis chlorophylla, Vulpicida pinastri, Parmeliopsis ambigua, Bryoria fuscescens; Table 1) were acidophytes.
The acidophyte richness (Fig. 2a) increased more strongly with distance from the city center than the nitrophyte richness (Fig. 2b). However, the response of both groups to distance diminished in the most recent 7-y period (Fig. 2). Interestingly, the richness of nitrophytes declined with distance from the city in the last 7-y period (P = 0.008) and was higher in the period 2013–2019 than in 1973–1979 across all distances, evidenced by non-overlapping 95% confidence intervals throughout the urban-suburban gradient (Fig. 2b). Consequently, the colonization of nitrophytes was much stronger than of acidophytes.
Fig. 2
The richness of a: Pseudevernion species and b: Xanthorion species along distance in km (log-scale) from Oslo city center for the first (1973–1979) and last (2013–2019) 7-year periods. Only sites within 10 km from the center were included. Linear regression lines (solid or hatched) and 95% confidence intervals (dotted lines) were given. Regression equations were as follow: a: Pseudevernion species richness 1973-79 = -2.796 + 9.970 x log(km). a: Pseudevernion species richness 2013-19 = 1.825 + 5.384 x log(km). b: Xanthorion species richness 1973-79 = -1.097 + 3.960 x log(km). b: Xanthorion species richness 2013-19 = 6.653–2.165 x log(km)
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The average species richness across urban-suburban sites (≤ 10 km) began to rise in the early 1970’s, with nitro- and acidophytes showing different rates of increase (Fig. 3a). An already substantial number of acidophytes at start of the period increased by a factor of 1.8 from 1973 to 2000, while the nitrophytes increased relatively faster by 3.6. The peak in species richness during 1998–2000 was followed by a noticeable temporal decline in 2001–2005 (Fig. 2a). Nitrophytes recovered from this stochastic decline after only 5 y, whereas the acidophytes required a longer period of 15–20 y to recover (Fig. 3a). Following this period of decline, the mean richness of acidophytes resumed its increase until the end of the study, while the increase in nitrophytes ended earlier (Fig. 3a). A 2-way ANOVA for the species richness across the seven periods in Fig. 3 and the two main lichen communities Pseudevernion and Xanthorion showed a clear period effect (P < 0.001), a weak community effect (P = 0.010), and a strong period x community-interaction (P < 0.001). The interaction implied mainly a faster increase in richness of Xanthorion than of Pseudevernion (Fig. 3a).
Fig. 3
a: The changes of mean species richness (± 1 standard error) across studied sites over a period of 47 years of total macrolichens (filled circles), Pseudevernion species (filled triangles), and Xanthorion species (open triangles). b: the total number of macrolichen species found in all studied sites in each period; in which c gives the numbers of studied sites. Means were averaged across each of the following periods: 1973–1975, 1976–1979, 1998–2000, 2001–2005, 2006–2010, 2011–2015, and 2016–2019
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The total number of species across urban-suburban sites increased from 26 in 1973–1975 to 34 in 1998–2000 and further to 44 species in 2016–2019 (Fig. 3b). Some of these differences between periods (Fig. 3b) could potentially be attributed to the varying numbers of studies conducted during each period (Fig. 3c). Interestingly, the rise in total species richness from 1973 to 1975 to 1977–1979 and from 2011 to 2015 to 2016–2019 (Fig. 3b) was noted despite a lower number of studies in the respective latter periods (Fig. 3c). In addition, the noticeable temporal decline in mean species richness in 2001–2005 (Fig. 3a) was not associated with a drop in total number of species (Fig. 3b). This decline was not an artifact of the few studies conducted during that period (Fig. 3c).
Discussion
Spatial variation in total macrolichen richness over ninety years
Thanks to the detailed lichen survey of Haugsjå (1930), trends in lichen richness can be prolonged to cover a time span of 90 years. The richness of macrolichens substantially declined from 1930 to the period 1973–1979 with a noticeable delay in the increase of species richness with distance, only reaching levels comparable to 1930 at the farthest distances (Fig. 4). After the 1970’s, lichens began to recolonize urban and suburban sites < 6–7 km from the city center. Within the inner 3 km, species richness in 2013–2019 was much higher than in 1930 (Fig. 4). The lichen struggle zone, as defined by Sernander (1926), extended about 5 km from the city center in 1930, expanding to approximately 7.5 km in the 1970’s (Fig. 4). The relatively low trendline peak in 2013–2019 (Fig. 4) implies that outer sections of the urban-suburban gradient are experiencing a permanent reduction in lichen richness, and that the normal lichen vegetation observed in 1930 no longer occurs in this part of the gradient. Due to incomplete knowledge of some species complexes in 1930, the total macrolichen diversity was then likely slightly underestimated. The increase in lichen richness with distance to the farthest rural sites for the years 1998–2019 (Fig. 1) suggests that the inner boundary of the normal zone is now at least 20 km from the city center.
Fig. 4
Trendlines for total macrolichen richness versus distance (≤ 10 km) from the city center of Oslo shown separately for 1930 (Haugsjå 1930), as well as for the first (1973–1979) and last 7-years period (2013–2019) in this study. Shaded bands show ± 1 standard error
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1930 to the 1970’s: increasing S-deposition and acidophyte richness
In 1930, when agriculture still occupied large areas in Oslo, the epiphytic macrolichen richness in the city was characterized by a diverse mixture of acidophytes and nitrophytes (Table 1). However, a significant change was observed from Haugsjå’s (1930) study to the 1970’s with a sharp decline in nitrophyte genera such as Anaptychia, Candelaria, Phaeophyscia, Physcia, Physconia, Polycauliona, Xanthoria. While such lichens earlier benefitted on dust from road soils and animal feces (Haugsjå 1930) that likely had high pH in this limestone-dominated region, the situation has changed after the urbanization of agricultural fields was completed in the 1970’s. The increasing acid rain towards 1970, which led to leaching of base cations from tree bark (Gauslaa 1995), might also have contributed to the decline of nitrophytes. Conversely, the frequency of acidophyte genera such as Bryoria, Hypogymnia, Parmeliopsis, Platismatia, Pseudevernia, Tuckermannopsis, and Vulpicida increased from 1930 to the 1970’s (Table 1). The major shift in epiphyte composition from 1930 to the 1970’s, which involved the replacement of Xanthorion by Pseudevernion, aligns with the documented increase in S-depositions and the related acidification before 1970 in Oslo (Gram 2005) and elsewhere (Lee 1998; Vestreng et al. 2007). Intriguingly, this shift towards acidophyte dominance occurred despite high N-deposition rates (Aas et al. 2017).
Lichens on acid bark (Gauslaa 1985) are associated with base cation-poor canopy throughfall and typically grow on conifers and birch (Gauslaa et al. 2021a). However, during the 1970’s, acidophytes occupied the bark of broadleaved deciduous trees (Øyseth and Aarvik 1980), a substratum that is not their usual habitat in unpolluted area due to its higher pH (Barkman 1958). This shift in phorophyte preference was likely boosted by increased air-borne acid rain originating from sources outside Norway (Menz and Seip 2004). The acid rain leached cations from the bark of broadleaved deciduous trees, effectively acidifying their bark to a pH level similar to that of conifers.
The period of peak S-deposition from 1963 to 1970 was not included in the study. However, lichen reinvasion and restoration of the lost buffer capacity in tree bark need time. Therefore, the lichen richness observed in 1973–1975 likely represents the period of most intense acidification. Prior to the early 1970’s, major reviews on lichens and air pollution did never mention lichen reinvasions (Ferry et al. 1973; Gilbert 1973; Hawksworth and Rose 1976). Nevertheless, the lichen richness in Oslo showed some recovery already between the initial and final period of the 1970’s.
1980–2019: low S-deposition and increasing nitrophyte richness
During the period of rapidly declining S-deposition (Gram 2005), epiphytic lichens began to colonize what was once a lichen desert in Oslo (Fig. 4), a phenomenon also observed elsewhere (e.g., Gilbert 1992). The lichen desert, which was present in 1930 (Haugsjå 1930) and in the 1970’s (Øyseth and Aarvik 1980), was no longer detectable despite the strong urbanization and doubling of the population. Better European control of S-emissions (Reis et al. 2012) facilitated the recovery of lichens. In the years marked by a slightly decreasing N-deposition (Grøntoft 2021), the urban increase in nitrophytes was substantial. For the first time, the richness of nitrophytes now significantly declines as the distance from the city increases. An upward trend in nitrophytic lichen communities has been documented in areas of intensive farming (De Bakker 1989) and on roadside city trees (Manninen et al. 2023).
Poor dispersal did probably not hinder recolonization, as many recorded lichens are early colonizers known as “zone skippers” during times of declining pollution (Gilbert 1992). Given the generally slow growth of lichens and their relatively rapid recolonization, it can be inferred that dispersal must have been effective from continuous forests on surrounding hills. The success of nitrophytic lichens in urban sites has weakened the previously strong gradient in lichen richness with distance from the city (Fig. 4), consistent with patterns observed in other cities (e.g., Gilbert 1992) where the disparity in SO2-concentration between the central and outer parts of the city has diminished (Laxen and Thompson 1987).
Are nitrophytic lichens nitrophilous species?
The unexpected surge in nitrophytes after 2001 during a period of weakly declining N-deposition (Aas et al. 2017; Grøntoft 2021), suggests that factors other than high N-emissions could be driving their proliferation. Firstly, the declining acidification probably facilitated the establishment of nitrophytes in vacant spaces when acidophytes that are intolerant to high N (Gaio-Oliveira et al. 2001, 2005; Manninen 2018) disappeared on broadleaved trees. Secondly, nitrophytes, which are reliant on higher pH (De Bakker 1989), often supplant acidophytes when bark pH increases (van Herk 1999; Herk 2001). In pristine conifer canopies, nitrophytes are strongly associated with locally high bark pH, high Ca, and low Mn concentrations in the canopy throughfall, but not with high N (Gauslaa et al. 2021a). Thirdly, while increased N-addition reduced growth rates of the nitrophyte Xanthoria parietina, the addition of other nutrients boosted its growth (Gauslaa et al. 2021a). This negative impact of N on growth was observed despite an N-induced increase in chlorophylls, an indicator of the photobiont population size (Palmqvist 2000). Increased N can even undermine the competitive advantage of nitrophytic lichens by causing an imbalance between their photo- and the mycobiont as the photobiont only is N-limited (Johansson et al. 2011; Palmqvist et al. 2017).
The term “nitrophyte” for Xanthorion species may warrant reconsideration. The process of reversed acidification of bark could have triggered the enhanced recolonization of nitrophytic species. It might be more appropriate to use neutral terminology, such as “eutrophs” in place of “nitrophytes”, and “oligotrophs” instead of “acidophytes”, as proposed by McCune and Geiser (2009). The terms Xanthorion and Pseudevernion are already neutral as they denote groups of co-occurring species.
It is, however, important to acknowledge that high NOX emissions from diesel vehicles (Tønnesen 2010 and correspondingly high N-concentration in roadside environments (Manninen et al. 2023) could locally surpass levels recorded by pollution monitoring sensors. A strong correlation between traffic noise and NOX in urban environments (Tenailleau et al. 2016) aligns with marked small-scale spatial variations in NOX, suggesting that N-concentration may have been locally underestimated. Given the ongoing strong move from fossil to electric vehicles and the anticipated reduction in NOX, it would be interesting to observe future shifts in lichen richness. In the context of Oslo, a city now bordered by the sea and forested hills with minimal agriculture, a future study could further examine the significance of N for nitrophytic lichens.
2001–2005: extreme autumnal rain can cause temporal lichen dieback
There were no abnormal pollution depositions observed on a national (Aas et al. 2017) or local level (Gram 2005; Grøntoft 2021) prior to and during the temporal drop in species richness starting in 2001. However, the completion of field work in 2000 was immediately followed by an unusually prolonged and continuous period of rainfall in the autumn. A rural study conducted outside Oslo attributed the decline of epiphytic lichens after 2000 to damage inflicted by incessant rain in the late autumn (Gauslaa 2023). This period of extreme wetness affected a large portion of southeastern Norway (Historic climatic data available at https://seklima.met.no/). Thus, the abrupt decrease in both acidophytes and nitrophytes after 2000 probably represented a rain-induced dieback in urban Oslo, a phenomenon explored in depth by Gauslaa (2023). Although the richness of macrolichen species in urban-suburban areas had rebounded by 2019 from this temporal dieback (Fig. 3), the recovery was less complete in rural areas with a larger species pool (Gauslaa 2023).
Conclusions
Epiphytic lichens, being spatially transferable pollution indicators (Delves et al. 2023), are highly sensitive to environmental changes. Their species richness exhibits significant responses to variation in pollutants within a few years and reacts even more rapidly to stochastic weather events. In a long-term perspective, successive pollution regimes and extreme weather can lead to severe degradation of the epiphytic vegetation. While considerable lichen damage and dieback have occurred, efficient control of S-emissions has facilitated regeneration, allowing a substantial portion of the previously existing epiphytic macrolichen vegetation to recolonize urban and suburban landscapes. Nonetheless, the lichen vegetation has not fully recovered. This incomplete recovery, coupled with persistent N-emissions, raises concerns for the future of species-rich acidophyte communities in populated landscapes.
Acknowledgements
I express my heartfelt gratitude to the numerous students for their fieldwork in the lichen course at the Norwegian University of Life Sciences over the years. Thanks also to Kari Balke Øyseth and Sissel Aarvik for their meticulous compilation of results from the years 1973-1979, and to the late Kåre A. Lye who initiated the lichen course in 1973 and guided it until 1975.
Declarations
Competing interests
The authors have no relevant financial or non-financial interests to disclose.
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