Do time-lagged extinctions and colonizations change the interpretation of buffer strip effectiveness? – a study of riparian bryophytes in the first decade after logging


Correspondence author. E-mail:


  1. There is a risk that short-term studies either underestimate disturbance effects because of time-lagged responses, including both time-lagged extinctions and colonizations, or overestimate them because of fast recovery.
  2. To evaluate the conservation effectiveness of tree group retention (in this case, buffer strips along streams), we studied the bryophyte community once prior to, and twice after logging, comparing one buffer and one clear-cut plot (0·1 ha) in each of 13 riparian sites. We asked whether time-lagged responses or recovery processes had dominated the period between two re-inventories, 2·5 and 10·5 years after logging, focusing both on the whole community and on species of conservation concern.
  3. Although there were examples of recovering species in both clear-cuts and buffer strips, the similarity in species composition to predisturbance conditions had decreased in the second re-inventory. Even if the buffer strips displayed more time-lagged colonizations and local extinctions over the later period compared to the clear-cuts, the overall species composition in the buffer strips was still significantly more similar to the prelogging conditions than the clear-cuts.
  4. The red-listed species had mostly declined during the first period, and the number of red-list species per plot (mostly species growing on dead wood) was rather stable at <20% of predisturbance levels in clear-cuts and <60% in buffer strips in the last re-inventory.
  5. Synthesis and applications. We show that most extinctions of red-list species occurred soon after disturbance and that the conclusions drawn from a study carried out 2·5 years after the disturbance did not change profoundly 8 years later. Although the species composition in the buffer strips continued to change over time, sensitive species survived much better in buffer strips than in clear-cuts, which supports the practice of retaining buffer strips for terrestrial species too. This knowledge should encourage managers to find ways of increasing the efficacy of this practice. One obvious measure could be to retain wider strips or implement other management practices that make the buffer strips less sensitive to wind, which will lead to higher tree retention to support a prelogging species composition.


Recently, there has been an increased interest in understanding extinction time-lags, also frequently termed ‘extinction debts’ usually in a context of human-imposed environmental change such as fragmentation or disturbance (cf. Kuussaari et al. 2009). An obvious and common reason for a time-lag in response to a changed environment is that the organism in focus is sessile and long lived, and as long as all individuals do not die immediately, the species can survive for some time even if it has ceased to reproduce and/or shows a reduced vitality (Vellend et al. 2006). Another kind of time-lag is the ‘immigration credit’, a term proposed for delayed colonization (Jackson & Sax 2010). When dispersal is a limiting factor, the actual colonization of a site might take some time, even if the site has become suitable, for example, through successional changes (Verheyen & Hermy 2004). In situations with large time-lagged responses, short-term studies might be of limited use for management decisions. In this study, we explore the extent to which time-lagged local extinctions and colonizations of boreal bryophyte communities change the conclusions regarding the efficiency of small forest patch retention (in this case, buffer strips along streams), by comparing studies carried out 2·5 and 10·5 years after logging.

Retention of live trees, either as single trees or in groups, on clear-cuts is one important measure in modern forestry aimed at mimicking natural disturbances, to some extent (Rosenvald & Lõhmus 2008; Gustafsson, Kouki & Sverdrup-Thygeson 2010). Studies of the distribution of fire refugia (e.g. Engelmark 1987) have been one basis for recommendations that retention patches preferably could be placed in wet and moist places (Vanha-Majamaa & Jalonen 2001). Buffer strips (retention of trees along watercourses) were initially used to protect aquatic values such as water quality and spawning sites for fish (Blinn & Kilgore 2001). Presently, however, buffer strip retention is a widespread and integrated part of overall tree retention strategies with the aim of ‘lifeboating’ (Lindenmayer & Franklin 2002) terrestrial species during clear-cut and young seral periods (Marczak et al. 2010), even if the motivation from a fire disturbance dynamics perspective has been shown not to be straightforward (Macdonald et al. 2004). More research is needed on the efficiency of retention patches in maintaining a prelogging forest species composition in the long run (Rosenvald & Lõhmus 2008; Perhans et al. 2009; Caners, Macdonald & Belland 2010).

Forests along streams are among the most species-rich places for bryophytes in the boreal landscape with between 40 and 120 species in an area of 1000 m2 (Hylander & Dynesius 2006), which is one reason why bryophytes have been investigated in relation to buffer strip management. An earlier study of the short-term effect (2·5 years) of both buffer strip retention and clear-cutting on the bryophyte community along small boreal streams showed that many species had suffered local extinctions, probably because of an unfavourable microclimate, while many other species had survived with rather high coverage (Hylander et al. 2005). Two and a half years is a short time and it is likely that a subsequent inventory would reveal large changes, both in terms of time-lagged and recovery responses, which could lead to different conclusions regarding the efficiency of buffer strip retention.

In this study, we describe how the vegetation characteristics have changed between the re-inventories at 2·5 and 10·5 years after the logging in 13 riparian clear-cuts and paired buffer strips. We compare the change in species composition in clear-cuts and buffer strips and determine whether time-lagged extinctions and colonizations have increased the overall dissimilarity to predisturbance conditions, or whether recovery processes have already resulted in a more late-successional forest-like community 10·5 years after the disturbance. We also study the red-list species separately, because the possibility of survival of such species in tree retention groups would decrease their dependency on large areas of natural forest habitat. Our study is important in the sense that we study the efficiency of green tree retention in a very species-rich group from the time before a disturbance, recently after the disturbance and >10 years after the logging event, whereas most other studies have been comparative studies or have covered shorter time periods (Rosenvald & Lõhmus 2008).

Materials and methods

Study System and Design

The study was conducted in boreal forests in central Sweden in the counties of Västernorrland and Jämtland (mid point of the area: 62°40′ N 16°05′, see Hylander et al. 2005). To a large extent, the area is a managed landscape with coniferous forests on acidic bedrock (for more details see Hylander et al. 2005). The forests have been managed for timber production for more than 150 years. Clear-cutting did not become the main logging method until the 1950s and 1960s, which means that the forests subject to clear-cutting, in most cases, have not regenerated after clear-cutting (Östlund, Zackrisson & Axelsson 1997).

In 1998, we selected 15 stands, intersected by a small stream (0·5–1 m wide), which were planned for clear-cutting by the forest company SCA (Svenska Cellulosa Aktiebolaget). All the selected stands were multi-aged mature stands with a minimum time of 86–146 years since regeneration, measured by coring an average-sized tree in each stand. Dominant or frequent tree species were Norway spruce Picea abies, Scots pine Pinus sylvestis and birch Betula spp. Most stands had signs of human management such as old stumps from selective logging. However, all sites had a mature forest bryophyte community with no signs of recent disturbance. In each stand, two 1000-m2 plots (20 by 50 m) were established along the stream with the stream crossing the middle of the short side of the plots. We also established one plot each in 10 reference stands not intended for clear-cutting. The references were similar to the treatment stands in terms of tree species composition and no signs of recent disturbance, but were probably on average older, and a few of them had considerably more dead wood. This was not considered a major problem because the main aim with the references was to account for changes in inventory efficiency over the years. In the winter of 1998/99, the 15 stands were clear-cut according to our instructions, leaving a buffer strip (10 m on each side of the stream) along one half of each study site and removing trees from the other half of the site (see Hylander et al. 2005). These sites were re-inventoried in 2001 and again in 2009. In 2009, only 13 sites remained for evaluation because the buffer strip had been removed from two sites. Nine of ten reference plots remained, as one site had been logged.

Data Collection

In each plot, we recorded all bryophyte species within five subplots (10 × 20 m), giving us a frequency measure of 0–5 for each species in each plot. Similar to the first re-inventory in 2001, we had the previous species list with us in the field in 2009. In the two-first inventories, only one person (KH) carried out the bryophyte inventory; however, in the most recent inventory, the work was divided between two skilled bryologists (KH and HW), and there was no bias in the recorder effort between the plot types. As before, the inventory of one plot (20 × 50 m) by one person took half a day. We collected many small samples of species that needed microscopy for a correct identification. However, to ensure comparable species lists among localities and among years, we grouped certain species groups together owing to, for example, many young individuals (see Appendix S2, Supporting information). We followed the checklist of Swedish bryophytes from 2006 (Hallingbäck, Hedenäs & Weibull 2006).

General environmental data were collected in the first year (1998), and variables that were related to changes probably caused by the logging were collected in the first re-inventory (2001) (Hylander et al. 2005). In 2009, we followed the same protocol for recoding basal area of different tree species, cover of different vegetation layers [canopy, shrubs (woody species <5 m), understorey and bryophytes (on ground)], number of uprootings, cover of mineral soil and amount of woody debris. We measured the length of the logs and counted the number of stumps in two size classes (10–20 and >20 cm in diameter at the narrowest point) and two decay classes (soft or hard wood). In 2009, we estimated the cover and height of regenerating trees of different species.


The statistical software r (version 2.10.1) (R Development Core Team 2010) was used for all analyses except where noted. Differences in selected vegetation characteristics among years and/or treatments were tested with pairwise Wilcoxon signed-rank test because of non-normal distributions of values.

The extent to which the species composition had changed between the two re-inventories was analysed by comparing the Bray–Curtis dissimilarity index among treatments (clear-cuts, buffer strips and references) and years. We calculated these values using both the presence/absence data and species frequency data. The differences between the two post-logging years and among treatments were evaluated by two linear mixed-effect models (using the package nlme, version 3.1-103): one that included clear-cuts and buffer strips (with site and plot as random factors; plot nested within site) and one that included buffer strips and references (with only plot as a random factor because the references were not in the same streams). Assumptions of normality and equal variances were checked by inspection of residual plots.

To understand whether the species compositional change was related to changes in the environmental variables (mostly disturbance intensity) in the buffer strips, we also calculated the Bray–Curtis dissimilarity index between 1998 and 2009 for the buffer strips. Thereafter, we conducted Pearson's correlation tests between these values and environmental variables that described the changes that had happened since the buffer strips were retained (i.e. number of tip-up mounds, change in canopy cover, change in basal area and cumulative length of fresh logs).

To investigate the differences in main trajectories in the response pattern among the treatment groups, including local colonizations and extinctions and time-lags in such responses, we defined nine different response types (Fig. 3a). We assigned all species with a difference in occupancy of 0, +1 or −1 between the three inventories as belonging to the stable category. If a species had changed its occupancy with ≥2 sites between two consecutive inventories, it was assigned to increasing or decreasing categories between these years. However, in the case of a changed occupancy of one site in the first period and another site in the second so that the total change was two sites, it was also classified as increasing or decreasing. Species with that response patterns (+1, +1 or −1, −1) were divided in equal proportions between the categories 2 and 3 (for increasing species) and 7 and 8 (for decreasing species) when plotted (See Fig. 3 and Appendix S2, Supporting information). Species displaying the following response patterns (−2, +1; +2, −1; −1, +2, or +1, −2) were classified as stable because of unclear responses and a total change of only one site between the first and last inventory. This analysis is not used to calculate the absolute number of species with different response types, but rather to illustrate the different response patterns within and among the treatments. The overall difference in frequency distribution of response types among the three treatments was compared with two Chi-squared tests: one between buffer strips and references and one between buffer strips and clear-cuts.

The mean proportions of species (compared to the pre-inventory data) that were initially lost from the plots (2001) and that were still absent (2009) were compared with the proportion that displayed a delayed local extinction in two linear mixed-effect models (in the same way as the Bray–Curtis dissimilarity index described earlier). In a similar manner, the contrasts were analysed between the proportions of species that recolonized vs. those that were still absent among the species lost immediately after logging. The proportion of new species in the plots was analysed in two ways: proportion of species with an early vs. a time-lagged increase (among the species that had colonized between 1998 and 2009) and the proportion of species that had colonized first but then became locally extinct vs. those that were still present in the second inventory among the newly colonized species recorded in the first re-inventory. All analyses were carried out as for the Bray–Curtis dissimilarity index described earlier.

The relationship between the change in occupancy and the change in local frequency (i.e. the mean number of segments [1–5] that a species was found in, including only the occupied plots) was analysed by Pearson's correlation tests for the two periods and for species increasing or decreasing in occupancy (≥2 plots) separately.

We summed the number of subplot records (number of 200 m2 records) of red-listed species for each plot and analysed both this value and the number of species per plot across treatments and years. Hylander et al. (2005) found that the number of records of red-listed species in 200-m2 subplots was similar to the number of easy delimited small patches (in general <0·01 m2) of these species in each plot. In the analyses, we omitted all red-list species found only in the references (mostly species on rock) and a somewhat doubtful species found only in the last inventory (Scapania cf. glaucocephala). Moreover, the Swedish red-list has been revised twice since the first inventory; however, we have only included those species named in the most recent red-list in this study (Gärdenfors 2010), which means that three red-listed species in Hylander et al. (2005) were omitted from the analyses. The analyses of the number of red-list species and the red-list subplot records consider the occurrence patterns of the following seven species: Anastrophylum hellerianum, Calypogeia suecica, Lophozia ascendens, Lophozia longiflora, Lophozia polaris, Scapania apiculata and Tayloria tenuis. The difference between the years and among the treatments in numbers of species and records was tested by Wilcoxon signed-rank test because of small numbers, but the results were displayed as percentage changes in the means for pedagogical reasons (change data could not be analysed because some cells would be divided by 0). However, we also analysed the percentage change from the prelogging situation omitting sites without records in the prelogging data (Buffer strips: N = 11, Clear-cuts: N = 8 and References: N = 7). With this approach, we could compare the change of species occupancy and species records among the treatment groups.


Vegetation Characteristics and Change

The mean and ranges of tree basal area and coverage of different vegetation layers were similar among the treatments prior to logging in 1998 (Table 1, for more details, see also Hylander et al. 2005). During the 8 years between the inventories in 2001 and 2009, there was a distinct increase in shrub cover both in clear-cuts (1–49%) and in buffer strips (6–30%) caused by the ongoing tree regeneration (most of the species were regenerating trees; Table 1). The maximum average height in regenerated vegetation was 3·2 m (Table 1). The opening of the canopy in the buffer strips has continued since 2001, with a mean cover change from 53% to 40% (= 0·031, Wilcoxon signed-rank test) accompanied by an increase in the total length of hard logs (>10 cm in diameter) from 137 to 205 m on average (< 0·001, Wilcoxon signed-rank test; Table 1). The mean bryophyte cover on the clear-cuts had started to recover from a mean of 30% in 2001 to 52% in 2009 (= 0·031, Wilcoxon signed-rank test). However, the variation was very large among plots ranging from sites with only around 10% cover to those almost totally covered with bryophytes [in the prelogging data, the cover was 70–100% (Hylander et al. 2005)].

Table 1. Vegetation characteristics (mean and range) in clear-cut plots (N = 13), buffer strip plots (N = 13) and reference plots (N = 9) in 1998 (prior to logging) and 2·5 and 10·5 years after logging of forests along small streams. Tree regeneration was not measured in 1998 and 2001, and soft logs were not measured in 2001
 Clear-cutsBuffer stripsReferences
  1. a

    Calculated without two incorrect values (probably measurement errors in either 2001 or 2009).

Basal area (m2×ha−1)35 (18–43)0·3 (0–2·3)1 (0–3·5)35 (18–56)21 (8–39)a18 (8–34)a27 (13–50)27 (16–44)
Logs (hard, >10 cm) (m)77 (7–153)17 (0–58)40 (10–104)76 (26–159)137 (0–363)205 (22–535)96 (26–218)44 (16–107)
Logs (soft, >20 cm) (m)6 (0–39)7 (0–24)5 (0–25)20 (4–55)13 (0–31)26 (13–37)
Tree cover (%)69 (50–90)1 (0–2)1 (0–10)66 (50–90)53 (5–80)40 (10–85)59 (30–90)64 (35–80
Shrub cover (%)15 (5–40)1 (0–4)49 (18–80)19 (1–50)6 (0–25)30 (1–70)13 (5–30)5 (1–20)
Tree regeneration (%)46 (18–81)29 (1–68)4 (0–15)
Height of regeneration (m)1·6 (0·8–3·2)1·6 (1·3–2·2)1·2 (0·5–1·7)
Broad-leaved regeneration (%)90 (73–99)62 (6–90)33 (0–77)
Understorey cover (%)82 (70–90)79 (79–92)90 (85–94)78 (40–90)76 (50–98)86 (65–93)68 (50–85)72 (50–90)
Bryophyte cover (%)83 (70–100)30 (10–60)52 (10–90)81 (40–100)71 (30–90)78 (10–94)84 (60–95)86 (60–94)

Species Compositional Changes

The overall bryophyte species composition was generally more dissimilar from prelogging composition in 2009 than in 2001 according to the linear mixed-effect models (F = 92, < 0·001 [clear-cuts and buffer strips]; F = 158, < 0·001 [buffer strips and references], Appendix S1, Supporting information; Fig. 1). The proportional change in species composition between 2001 and 2009 was higher in the buffer strips compared to clear-cuts (53% vs. 20%, t = 2·6, = 0·022, paired t-test). However, buffer strips had still experienced significantly lower compositional change than clear-cuts and still retained its intermediate position between clear-cuts and references (Fig. 1, Appendix S1, Supporting information). The pattern was similar if the presence/absence data or the frequency data (not shown) was used to calculate the Bray–Curtis dissimilarity index. Number of tip-up mounds in buffer strips (which could be regarded as a proxy for amount of disturbance after the retention of the strips) correlated significantly with bryophyte compositional change (Bray–Curtis dissimilarity index between 1998 and 2009, r = 0·59, = 0·035, Pearson's correlation test, Fig. 2). Similar correlations were also obtained for dissimilarity vs. change in canopy cover, change in basal area and length of fresh logs that all describe the same gradient in disturbance over the investigated years.

Figure 1.

Species compositional distances between years in clear-cuts, buffer strips and references based on Bray–Curtis dissimilarity index using presence/absence data. Higher values denote larger changes in species composition in a plot between two measurements. For statistical interpretation, see 'Results' section and Appendix S1 (Supporting information). Error bars = SE.

Figure 2.

Dissimilarity (Bray–Curtis) between species composition in each buffer strip plot in 1998 and 2009 in relation to number of tip-up mounds in the same plots (0·1 ha). A linear trend line is added for clarity (Pearson's correlation test, r = 0·59, = 0·035).

Species Response Types

In total, 280 species (taxa) were found in the study (Appendix S2, Supporting information). In clear-cuts and buffer strips, there were species displaying many different trajectories over time (as defined by our classification of species into response categories [see 'Materials and methods' section]; Fig. 3, Appendix S2, Supporting information). The overall frequency distribution of the different response types differed significantly between buffer strips and references (χ2 = 35·1, < 0·001) as well as between buffer strips and clear-cuts (χ2 = 15·6, < 0·048).

Figure 3.

Response trajectories of species between 1998 (prelogging) to 2001 (2·5 years after logging) and 2009 (10·5 years after logging). (a) Illustrates the different possible trajectories and b–d shows the trajectories in (b) clear-cuts, (c) buffer strips and (d) references. The trajectories (in a) represent as follows: (1) steady increase, (2) initial increase and then stabilizing, (3) time-lagged increase, (4) initial increase followed by a return to predisturbance levels, (5) stable, (6) initial decline followed by an increase towards predisturbance levels, (7) time-lagged decline, (8) initial decline and then stabilizing and (9) continuing decline. In (b–d), the width of the lines represents the proportion of species displaying the particular response trajectory. The inclusion of a species in a certain group is defined by its change in occupancy between the inventory years (See 'Materials and methods' section and Appendix S2, Supporting information). Number of species and the relative proportion in each group are given in the figures.

Proportion of Species Going Extinct and Colonizing at the Plot Level

The proportion of species that had gone locally extinct per plot in 2009 compared to the prelogging situation was 12% in the buffer strips, which was significantly less than in the clear-cuts (18%, = 0·0096, paired t-test), but significantly more than in the references (7%, = 0·0029, t-test; see also results from linear mixed-effect models, Appendix S1, Supporting information). The proportion of the species that was lost during the second period (compared to the whole period), thus, displaying a time-lagged local extinction was 56% in the buffer strips compared to 40% in the clear-cuts (significant interaction in the linear mixed-effect model, F = 6·3, = 0·019, Appendix S1, Supporting information; Fig. 4a). There was a higher proportion of the species in the buffer strips than in the references that had colonized the plots between the prelogging and the first re-inventory as well as between the first and second re-inventory [< 0·001, Appendix S1 (Supporting information), Fig. 4b], while the levels in the buffer strips and clear-cut plots were similar (Fig. 4b). On average, only a few of the species had first colonized and then become locally extinct (Fig. 4b).

Figure 4.

Mean proportions of (a) species going locally extinct from plots and (b) species colonizing plots. The following contrasts were analysed statistically (see 'Results' and Appendix S1, Supporting information): In (a) the proportion of species that became locally extinct immediately (among the species that still were not present) vs. the proportion that declined after the first re-inventory and the proportion of species that recolonized vs. those that became locally extinct and not did recolonize. In (b): the proportion of species that colonized immediately vs. the proportion that colonized after the first re-inventory and the proportion of species that colonized directly and persisted vs. the proportion that colonized directly but later on became locally extinct.

Change in Occupancy vs. Change in Local Frequency

There were no significant correlations between species with a decreasing occupancy and their change in local frequency at the sites they were still present in (Appendix S3, Supporting information). For species with an increasing occupancy, there was also a pattern of an increasing local frequency in the first period in the clear-cuts (r = 0·44, = 0·011, Pearson's correlation test) and in the second period in the buffer strips (r = 0·40, = 0·013, Pearson's correlation test, Appendix S3, Supporting information).

Red-Listed Species

Very few red-list species survived the clear-cut treatment, and the mean number of species as well as subplot records was still only around 20% of the pre-inventory numbers 8 years after the first re-inventory (Fig. 5, P > 0·85 between 2001 and 2009, Wilcoxon signed-rank tests). The mean number of red-list species and subplot records in buffer strips did not change significantly between the re-inventories (> 0·78, Wilcoxon signed-rank tests), and the means were around 60% of the prelogging levels (Fig. 5). The level of protection of red-list species and subplot records in the buffer strips might be considered to be even lower than the absolute values indicate, if the prelogging underestimation (seen by the increasing bars in the references) is taken into account. The proportion of both red-list species and red-list records in the last re-inventory compared to the prelogging data was significantly higher in buffer strips compared to clear-cuts but significantly lower in buffer strips compared to references (< 0·05, Wilcoxon signed-rank test).

Figure 5.

Change in (a) mean number of red-list species records and (b) mean number of red-list species between 1998 (before logging) and 2001 (2·5 years after logging) and 2009 (10·5 years after logging), respectively. A number of red-list species only found in the references are omitted from the analysis (see 'Materials and methods'). There were no statistically significant change between 2001 and 2009 in any of the groups (See 'Results' section).


There has been an increased awareness that time-lags are common in nature and must be understood to correctly predict the outcome of, for example, fragmentation or restoration (Vellend et al. 2006; Jackson & Sax 2010). Green tree retention is a fairly new practice and there is a great need of studies evaluating the effects of this practice over longer time periods, because one of the main aims is to preserve a late-successional forest-like environment and its species throughout a forest regeneration cycle (Gustafsson, Kouki & Sverdrup-Thygeson 2010). We found that the overall species composition in buffer strips had continued to diverge from prelogging conditions between the two re-inventories 2·5 and 10·5 years after the logging, indicating some time-lagged responses, even if the bottle neck in microclimate and substrate availability seemed to have passed for some species, which had regained their previous frequencies (cf. Lõhmus & Lõhmus 2010). The overall pattern was a greater time-lag in both colonization and extinctions in buffer strips compared to clear-cuts, but generally also more delays in colonizations than in extinctions in both clear-cuts and buffer strips (Figs 3 and 4).

Buffer Strip Efficiency

Forests along small streams can be very important habitats for bryophytes sensitive to forestry in boreal forests, and this study shows that the retention of even narrow buffer strips can make a difference. The fear that there was a large extinction debt in the first re-inventory and that subsequent dry years had extirpated the last populations of sensitive species from the buffer strips was largely unwarranted. Even if our results showed more delayed extinctions at the plot level in buffer strips (Fig. 4a), it also showed that the local extinction of red-list species (at both investigated scales) had already occurred during the first period (Fig. 5). The judgement from the first re-inventory that buffer strips protected around 50–60% of the subplot records of the red-listed species was, thus, also still valid after another 8 years even if some species certainly were more difficult to protect in buffer strips compared to other species (Appendix S2, Supporting information). Another study comparing retention groups over a 6-year period showed declining populations of red-list and indicator bryophytes (Perhans et al. 2009), but it is difficult from that study to judge whether it was an immediate or a time-lagged decline. Thus, we conclude that short-term studies on disturbance effects on bryophytes can be informative in terms of which species decline or become locally extinct due to changed habitat conditions, but we acknowledge that our study is also comparatively short term in the context of forest succession. Our results also support the findings of fast stabilization of epiphytic bryophyte and lichen communities found after clear-cut disturbance in Estonia (Lõhmus & Lõhmus 2010), but contrasts to studies on, for example, vascular plants where the extinction process tends to be slower (Vellend et al. 2006). Knowing that even narrow buffer strips can be valuable for the most sensitive species should encourage managers to find ways of increasing the efficacy of this practice. Microclimatic edge effects have been shown to affect bryophytes 40–50 m from the edge in mesic boreal forests (Hylander 2005). Even if the edge effects might be less severe in concave and moist landscape settings, this indicates that considerably wider buffer strips would be needed in order for most of the sensitive species to thrive. Wider buffer strips would also be less sensitive to blowdowns and thus change less in terms of total species composition (cf. Fig. 2). The loss of host trees in retention groups due to wind-throw and other mortality is also considered an important cause of local extinction over time for epiphytic organisms such as lichens, diminishing the ‘lifeboating’ value of such retention groups (Lõhmus & Lõhmus 2010), which supports the idea of retaining buffer strips wider than 10 m on each side of watercourses.

Buffer Strips as Lifeboats over Time

Based on the assumption that each metapopulation has a capacity (Hanski & Ovaskainen 2002), most studies on extinction debt at a landscape scale assume that a decrease in connectivity might lead to future extinctions (Bulman et al. 2007). Therefore, we cannot assume that the extinction debt at that scale already has been paid off after 10·5 years, which from a forest ecology perspective, should also be considered to be a short time. A future decline of remaining red-list species might be possible given the small area of old forests in the landscape. A slow re-colonization rate has also been highlighted by other authors as a severe problem for the conservation capacity of retention groups (Lõhmus & Lõhmus 2010). However, because a large amount of downed wood was created in the buffer strips during these years, another possible scenario might be that these substrates (if they are left by the landowners) at some point will be colonized by specialist species (downed wood was still too hard during this study) and that large reproducing populations will build up, especially if they have survived in small pockets somewhere in the buffer strips. Even if the most wind disturbed buffer strips provided the least protection for the prelogging bryophyte communities, they might still be valuable both in the long- and short term for many other organisms given the large amount of downed wood (Jonsson, Kruys & Ranius 2005), and as such function as temporal ‘lifeboats’ at a landscape level. More research is needed on how different organisms make use of different time windows in retention groups and on single retained trees, and to what extent acclimatization to a lighter and drier environment occurs (cf. Jairus, Lõhmus & Lõhmus 2009).

Time-Lagged Colonizations

A greater number of species colonized the clear-cuts than the buffer strips during the first two and a half years after the logging (Fig. 3, Hylander et al. 2005), even if the difference was not large at the plot level (Fig. 4b). Many of these were species commonly found in wet and light environments such as mires, but a large group was small pioneers on mineral soil (mesic, moist or wet). There has been a steady new production of exposed mineral soil in the buffer strips, caused by fallen trees, which can explain the later increase in local frequency of such species there (Fig. 3, Appendix S3, Supporting information). Most pioneer species that colonized the clear-cut soon after the disturbance event continued to increase their occupancy and frequency or at least stay at the same level even if there was little production of new exposed soil there [but see Leptobryum pyriforme for an exception (Appendix S2, Supporting information)]. Invading species after disturbances can be a conservation problem. However, in this system, all these species are confined to exposed substrates and are eventually out-competed by longer-lived species. In a study of stream-side forests 30–50 years after clear-cutting, these pioneer species had disappeared to a large extent (Dynesius & Hylander 2007).

The Importance of References

The mean species number in the references increased from 90 (1998) via 94 (2001) to 101 (2009). We believe that this is mainly a result of an increased efficiency in finding species because of previous knowledge of the content of the plot and an inevitable increased inventory skill over the years. The fact that many species showed an increasing trend in the references (both at subplot and plot level) supports this interpretation (see Appendix S2, Supporting information). Hence, we believe that there was a prelogging underestimation in all plots, which we have handled by comparing the various results in the treatments with the results in the references (the standard usage of references). Only 9 of the 31 increasing species in the references also belonged to the increasing species in buffer strips and/or clear-cuts with the inclusion criteria used (Appendix S2, Supporting information). However, the increases for Cephaloziella sp., Pohlia bulbifera and Polytrichastum longisetum, were much larger in the treatment plots (mean increased occupancy of 7·7 and 5) compared to the reference plots (2·7) indicating that these species were correctly placed in this category. Our results show that it can be difficult to make time series over large areas using small organisms requiring special expertise and highlights the necessity of using references in such cases.


We demonstrate that buffer strips can protect (to a certain extent) species that become extinct at clear-cutting and that this effect is not eroded over a decade. This would encourage managers to find ways of increasing the efficiency of their retention groups. One measure could be, for example, to increase the buffer width at selected places to decrease the direct (e.g. wind-throws) and indirect (e.g. microclimatic edge effects) disturbance effects, thus protecting a larger proportion of sensitive species (cf. Aubry, Halpern & Peterson 2009; Perhans et al. 2009). The question is open, whether or not the large amount of dead wood in the buffer strips will be colonized by bryophyte species of conservation focus when it becomes suitable (which might take another decade). This will depend on whether the focal species still occurs in situ and/or to what extent the density of these species in the landscape is high enough to provide propagules before the dead wood decomposes. This kind of role for retention groups is even less studied than the direct survival of species after the logging event and deserves much attention.


Thanks to Ola Svensson for field assistance in 2009. We want to thank the company SCA for the help in finding field sites. Many more persons have been helpful during both the pre- and postlogging inventories and species verifications in 1998 and 2001 (See Hylander et al. 2005). Mats Dynesius, Jörgen Rudolphi and three anonymous reviewers provided valuable comments on the manuscript. The re-inventory was supported by a grant from ‘Stiftelsen Carl-Fredrik von Horns fond’ [to KH].