Proteaceae juvenile periods and post-fire recruitment as indicators of minimum fire return interval in eastern coastal fynbos

Abstract

Question

Fire management practices aimed at biodiversity conservation are often in conflict with hazard reduction requirements. Particularly in protected areas where such conflict of interest exists, the question is asked: what are the ecological thresholds within which fire should be managed?

Location

Montane proteoid fynbos shrublands, eastern coastal part of the Cape Floral Kingdom; coastal slopes of the Tsitsikamma and Outeniqua Mountains, South Africa.

Methods

Estimates of optimal fire frequencies are often based on the relationship between plant age and the rate of seed accumulation of the slowest-maturing species. We established juvenile periods and recruitment success (measured as the ratio of post-fire recruits to the pre-burn population) after fires at different intervals, of serotinous, reseeding shrubs in the Proteaceae. From this we estimated minimum fire return intervals (FRIs) that would allow for their persistence in eastern coastal fynbos shrublands.

Results

Juvenile periods ranged from 4 to 9 yr, which are comparable to ecologically similar proteoids elsewhere in fynbos and SE and SW Australian shrublands. There was large variability among sites and within species in the degree of flowering at given plant ages, some of which seemed related to the moisture regime. There were no consistent differences among species in their rate of maturation. Post-fire recruitment success was near zero following a fire in 5-yr-old vegetation, always above replacement levels following fires in ≥7-yr-old vegetation, and at a maximum in old (38 yr) fynbos. There was considerable variation in post-fire recruitment success for any particular FRI, species or site. The lack of a significant relationship between recruitment success and pre-fire vegetation age, suggests that once a critical post-fire age is attained, factors other than seed abundance affect recruitment success.

Conclusions

From an ecological perspective, our findings imply a minimum FRI of 9 yr for eastern coastal fynbos. This is not intended to prescribe rigid management of fire according to a fixed rotation and does not negate the need to consider site- or species-specific requirements. Instead it provides a lower threshold for a range of acceptable FRIs below which a significant decline of species populations is predicted.

Nomenclature

Rebelo

(2001)

Introduction

Ecosystem integrity can be compromised by disturbance regimes that are outside of the historical range of variability to which the ecosystem's component species are adapted (DellaSala et al. 2004). Fire is such a disturbance regime, and empirical evidence indicates that variation in fire return intervals (FRIs) can affect species abundance and floristic composition in mediterranean-type ecosystems and other temperate shrublands (van Wilgen 1981; Haidinger & Keeley 1993; Morrison et al. 1995; Lloret et al. 2003).

Life-history attributes that render species vulnerable to certain disturbance regimes, especially variations in the interval between fires, are well recognized (Noble & Slatyer 1980; van Wilgen et al. 1992; Bond & Wilgen 1996; Bradstock & Kenny 2003). Obligate reseeders are particularly prone to population decline under increased fire frequency when the interval between fires is shorter than the time taken to accumulate a seed bank adequate for post-fire regeneration (Morrison et al. 1995; Knox & Morrison 2005). Obligate reseeders with canopy-stored (serotinous; Lamont et al. 1991a) seed banks are most susceptible to short FRIs because their whole seed bank is released post-fire (Noble & Slatyer 1980; Bradstock & Kenny 2003). Estimates of ecologically appropriate FRIs for a site are thus often based, at the lower end of the range, on the age of maturation and rate of seed accumulation of the slowest-maturing obligate reseeder species, and at the higher end, on the life span of the shortest-lived reseeder species that does not accumulate soil seed banks (Bond 1980; Enright et al. 1996; Gill & McCarthy 1998). At both ends, the most sensitive plant guild is the slow-maturing, short-lived, serotinous, reseeding shrubs (Noble & Slatyer's (1980) CI species) in South African and Australian temperate shrublands, typically of the Proteaceae family (hereafter termed ‘proteoids’) (Cowling 1984; Kruger 1984; Le Maitre & Midgley 1992; Enright et al. 1996, 1998; Gill & McCarthy 1998; Bradstock & Kenny 2003).

Fire return intervals equal to twice the primary juvenile period (time elapsed between germination and flowering) were shown to be necessary to allow for adequate seed production in western Australian jarrah forest (Bell 2001). Similar methods of determining appropriate minimum fire intervals have also been suggested for eastern Australian shrublands (Gill & McCarthy 1998) and fynbos shrublands of the Cape Floral Kingdom (CFK) (Kruger & Lamb 1978). A rule of thumb used in the fynbos is that half of the individuals in a population of the slowest-maturing proteoid species should have flowered in at least three successive seasons before the area may be burned (Kruger & Lamb 1978; Kruger 1982).

There is a need to incorporate ecological thresholds for FRIs into fire management systems (Bradstock & Kenny 2003; van Wilgen et al. 2011), particularly in protected areas where the primary goal is biodiversity conservation. Fire management practices aimed at ecological objectives are, however, often in conflict with hazard reduction requirements (Morrison et al. 1996; Driscoll et al. 2010; van Wilgen et al. 2012). Simple management compromises intended to reconcile conflicting objectives may ultimately not achieve either hazard reduction or biodiversity conservation (Morrison et al. 1996). In the Garden Route coastal mountains (GRCMs), conflict of interest in terms of desired FRIs is a matter that remains to be addressed (Kraaij et al. 2011). First, the landscape comprises fynbos, of which large tracts occur within the newly established Garden Route National Park (GRNP), interspersed with plantations of fire-sensitive alien pine trees. As a result, there is significant pressure from the plantation industry to reduce wildfire hazard by burning fynbos, in places, at short intervals (Kraaij et al. 2011). Second, the GRCMs occur within the eastern coastal part of the CFK where fynbos fire ecology in general, and thresholds for FRIs in particular, are inadequately understood (van Wilgen 2009). Third, high levels of invasion of fynbos by alien trees (most of which originate from the timber plantations) require integration of alien plant control considerations with fire management measures (Richardson 1998). As the GRNP has only recently been established, it is important to influence the development of management policy and practice in the interests of biodiversity conservation (Kraaij et al. 2011).

In this study we established juvenile periods and recruitment success (Noble & Slatyer 1980) in proteoids after fires at different intervals, and used this information to estimate minimum FRIs (van Wilgen et al. 2011) that would allow for their persistence in eastern coastal fynbos.

Methods

Study area and species

We focussed on the GRCMs (33.80° S 22.59° E–34.01° S 24.26° E) that occur along the Cape south coast of South Africa. The area includes the southern slopes of the Outeniqua (east of the Touw River) and Tsitsikamma Mountains. Within the study area, we distinguished between the Outeniqua and Tsitsikamma regions, to the west and east of the Keurbooms River, respectively (Fig. 1). The climate of the GRCMs is relatively temperate, owing to maritime influence. In contrast to the strictly mediterranean climate in the western part of the CFK, rain falls more evenly throughout the year (Schulze 1965). Rainfall peaks in spring and autumn in the GRCMs, and winter months are the driest. The proportion of summer rain increases eastwards (Schulze 1965; Tyson & Preston-Whyte 2000). Mean annual rainfall increases eastwards, from 820 mm in the Outeniqua to 1079 mm in the Tsitsikamma Mountains (Bond 1981; Southwood 1984). Hot and desiccating winds that flow from the interior (known locally as bergwinds) occur most frequently during autumn and winter, when they increase the likelihood of fires (Geldenhuys 1994; Southey 2009). The weather conditions suitable for fires dominate in the dry summer months in the west of the CFK, but become progressively less seasonal towards the east (T. Kraaij, unpublished data).

Figure 1. Locations of sites where three different types of surveys were done of proteoids, i.e. recurrent surveys of growth and flowering status, one-off surveys of flowering status, and post-fire recruitment success. The study area is divided into the Outeniqua and Tsitsikamma regions, to the west and east of the Keurbooms River, respectively.

The fire-prone vegetation within the study area includes ca. 100 000 ha of mountain fynbos shrublands (Rebelo et al. 2006) and 47 000 ha of commercial pine plantations (Kraaij et al. 2011). Fire-resistant indigenous forests occur largely on the coastal plateau along the foothills of the mountains, as well as in fire refugia such as gorges, scree slopes and bergwind shadows in the mountains (Geldenhuys 1994). Large parts of the GRCMs are remote, and natural fires accounted for 60% of the area burned during the past century (T. Kraaij, unpublished data). Fynbos comprises fire-dependent (Kruger & Bigalke 1984) evergreen shrublands with evergreen graminoid understoreys (Rebelo et al. 2006) that burn at median intervals of 10–21 yr elsewhere in the CFK (van Wilgen et al. 2010) and at 8–26 yr locally (T. Kraaij, unpublished data). Overstorey proteoids common to the fynbos of the region are Protea neriifolia, P. mundii and Leucadendron eucalyptifolium, while L. uliginosum and P. eximia occur less commonly. In the eastern part of the CFK, there is a shift from winter–spring flowering to summer–autumn flowering compared to the west (Pierce & Cowling 1984; Johnson 1993; Heelemann et al. 2008), both across and within lineages.

Study design

Data collection was non-manipulative, relying on the sampling of areas in the field with known recent histories of fire occurrence. We undertook three types of surveys (Fig. 1) of proteoids:

1. One-off surveys to determine juvenile periods of populations where the date of the last fire, and thus plant age (i.e. time since last fire), was known (range 4–11 yr). We randomly selected 50–250 individuals per species per site by means of the wandering quarter method (Catana 1963). For each individual, we recorded the number of times (seasons) it has flowered. We located six survey sites within the Tsitsikamma region (Fig. 1), two of which were sampled twice, 2 yr apart. Most sites had more than one species of interest and some sites offered different habitat types related to aspect and slope. In total, 29 site–species replicates were sampled and 2789 plants.
2. Recurrent surveys of post-fire growth and flowering of permanently marked plants at sequential plant ages at two sites. The first site was on a dry west-facing slope at Keurbooms Nature Reserve (200 m a.s.l.) after a prescribed fire (223 ha) in April 1996, where 100 seedlings of P. neriifolia were marked along ten transects (spaced 2-m apart) in 2006. Plant height from ground level and the number of flowerheads of the current season were recorded annually until 2011. The second site was on a moist south-facing slope at De Vasselot Nature Reserve (280 m a.s.l.) after a lightning fire (29 855 ha) in November 2005, where 100 seedlings each of P. mundii and L. eucalyptifolium were marked along six transects (spaced 10-m apart) in 2008. Plant height from ground level and the number of flowerheads of the current season were recorded at approximately annual intervals until 2011. The two sites are situated 15-km apart and both are ca. 5 km from the coast.
3. Surveys of recruitment success within 4 yr post-fire (mean ± SD = 1.8 ± 0.5 yr), where populations occurred in areas where the pre-fire vegetation age was known (range 5–38 yr). We counted within belt transects (2 m × 30 m) the number of proteoid seedlings (post-fire recruits) in relation to the number of burned parents, using the non-destructive method of Bond et al. (1984) as modified by Husted et al. (2007). We located 12 survey sites throughout the study area (Fig. 1), which provided 22 site–species replicates. The mean (±SD) number of transects per site–species replicate was 5.0 ± 2.8.

Data analysis

Analyses of the three types of survey data outlined above were done as follows:

1. We calculated for each site–species replicate, the proportion of the population that had flowered at least once, twice and three times or more at the time of surveying (at known plant ages).
2. We calculated, for each plant age sampled and for the species and sites separately, the mean height of survivors, the proportion of the original marked population flowering, the proportion of the population surviving, and the mean number of flowerheads per flowering plant. We explored the relationship between plant height and the number of flowerheads per flowering plant at the time of the last survey (2011) for each of the respective species by means of least-squares regression.
3. We calculated the seedling–parent ratio for each site–species replicate and applied a square root transformation to conform to the assumption of normal distribution (Kolmogorov–Smirnov, P = 0.91). We assessed the relationship between seedling–parent ratio and pre-fire vegetation age for all species combined using least-squares regression. Season of burn, age of the vegetation at the time of surveying, and pre-fire parent density (Bond 1984; Bond et al. 1984; Le Maitre 1988a,b; Heelemann et al. 2008) may additionally affect recruitment success. However, in our sample, the seedling–parent ratio did not differ significantly between cool-season (June–November; 10.2 ± 5.7, mean ± SD, n = 6) and warm-season (December–May; 9.4 ± 9.5, n = 13) fires (cf. Heelemann et al. 2008), neither was it influenced by the (post-fire) age of the vegetation at the time of surveying (F_1,20 = 0.72, P = 0.41, R² = 0.03), nor by parent density (F_1,20 = 0.03, P = 0.87, R² = 0.01). We therefore did not account for these factors in our analysis of the effects of pre-fire vegetation age on recruitment success.

Results

Juvenile period

Juvenile periods of proteoids ranged from 4 to 9 yr (Table 1). There was considerable variation in the degree of flowering at any given plant age among sites and within species (Fig. 2). There were no consistent differences among the three most common study species in their rate of maturation. Across sites and species, <40% of individuals had flowered once by the age of 4 yr, while more than half had flowered once by the age of 5 yr. At 11 yr of age, 90–100% of individuals had flowered at least once, more than half had flowered at least twice, while 28–83% had flowered at least three times.

Figure 2. Proportions of populations of Leucadendron eucalyptifolium (Le), Protea mundii (Pm), Protea neriifolia (Pn) and Protea eximia (Pe) that have flowered at least (a) once, (b) twice, and (c) three times at different post-fire vegetation ages.

Table 1. Primary juvenile periods (yr) of reseeding proteoids in the Tsitsikamma region, the Outeniqua Nature Reserve (immediately west of this study's Outeniqua region) and the western part of the Cape Floral Kingdom (CFK). The post-fire age (yr) at which 50% of the individuals in proteoid populations flowered once and three times are also shown

	Eastern-CFK		Western-CFK
	Tsitsikamma*	Outeniqua†
Source: ^*Current study; ^†CapeNature (unpublished data); ^‡Le Maitre (1992); ^§Kruger & Bigalke (1984).
Juvenile periods
Protea neriifolia	4–9	4–6	4–5‡
P. mundii	4		7§
P. repens		≥4	3§
P. lorifolia		5
P. eximia	4		3§
50% of population flowered once
P. neriifolia	5–6	8–11	8‡
P. mundii	6
P. repens		10
P. lorifolia		7
50% of population flowered 3×
P. neriifolia	11	8–14	12‡
P. mundii	11
P. repens		14–15
P. lorifolia		>13

Mortality of permanently marked plants was low, with 93% of L. eucalyptifolium and 98% of P. mundii surviving 4 yr post-marking and 6 yr post-fire, and 95% of P. neriifolia surviving 6 yr post-marking and 15 yr post-fire. The number of flowerheads per flowering plant was positively related to plant height in L. eucalyptifolium (F_1,30 = 12.7, P = 0.001, R² = 0.30) and P. mundii (F_1,24 = 7.6, P = 0.01, R² = 0.24), but not in P. neriifolia (F_1,28 = 2.9, P = 0.10, R² = 0.09) at the time of the last survey (2011). The number of flowerheads of the current season per flowering plant did not vary widely within or among years (apart from L. eucalyptifolium where limited data restricted inference) and did not appear to increase with time during the periods surveyed (Fig. 3). The post-fire age at which populations attained a particular height or degree of flowering differed substantially between the two sites, but not between P. mundii and L. eucalyptifolium on the moist southern slope. A few individuals of P. mundii but none of L. eucalyptifolium started flowering at 4 yr of age, while ca. 30% of the populations flowered by 6 yr of age. At the drier site, few P. neriifolia flowered by 9–10 yr of age, while only ca. 30% of the population flowered by 13–15 yr of age.

Figure 3. (a) The proportion of the population flowering, the above-ground plant height (mean ± SE) and (b) the number of current-season flowerheads per flowering plant (mean ± SE) at sequential post-fire vegetation ages of Protea mundii (Pm), Protea neriifolia (Pn) and Leucadendron eucalyptifolium (Le). Pm and Le were measured at 1–6 yr post-fire on a relatively moist, south-facing slope, and Pn at 9–15 yr post-fire on a relatively dry, west-facing slope.

Post-fire recruitment

Seedling–parent ratios (of all species collectively) were not significantly correlated with pre-fire vegetation ages (F_1,20 = 2.87, P = 0.11, R² = 0.13), and ranged from 0 to 35.7 (Fig. 4). Post-fire recruits of P. neriifolia were absent from the site that burned at 5 yr of age, while recruitment was below replacement levels in L. eucalyptifolium and P. mundii (seedling–parent ratios of 0.05 and 0.11, respectively). Juvenile period surveys done at this particular site 1 yr prior to the fire showed that <2% of the individuals of P. neriifolia and L. eucalyptifolium flowered once, while 25% of the individuals of P. mundii flowered once at the post-fire vegetation age of 4 yr. Overall, seedling–parent ratios varied widely (coefficient of variation 105%) within and among species as well as within pre-fire vegetation age classes (Fig. 4).

Figure 4. Post-fire recruitment in relation to the age at which the vegetation burned. Recruitment is expressed as the ratio of the number of seedlings to burned parents, and shown for four proteoid species, i.e. Leucadendron eucalyptifolium (Le), L. uliginosum (Lu), Protea mundii (Pm) and P. neriifolia (Pn).

Discussion

Juvenile period

Juvenile periods of proteoids ranged from 4 to 9 yr in our study, which are comparable to those at Outeniqua Nature Reserve, immediately west of the study area, those reported for the western part of the CFK (Table 1; CapeNature unpublished data; Kruger & Bigalke 1984; Le Maitre 1992), and those of proteoids in SE and SW Australia (Cowling et al. 1987; Richardson et al. 1990; Lamont et al. 1991b; Enright et al. 1996, 1998; Gill & McCarthy 1998; Bell 2001; Bradstock & Kenny 2003). However, there is usually a lag between the age at which the first plants flower and that at which the majority flower (Kruger & Bigalke 1984). Early inflorescences furthermore do not necessarily produce fertile cones (Enright et al. 1996). In western Australian scrub heath, a detailed study of the canopy seed bank dynamics of Banksia hookeriana showed that even though the species started flowering at 3–4 yr of age (similar to our findings), FRIs optimizing the likelihood of successful recruitment were estimated at 15–18 yr (Enright et al. 1996).

In North American temperate serotinous pine forests, under regimes of frequent disturbance, maturation ages optimizing plant fitness were estimated to be 0.4 times the average disturbance interval (Clarke 1991). Median FRIs recorded in the study area since 1980 varied considerably (8.3–26.2 yr) depending on the method of estimation (T. Kraaij, unpublished data), and predict optimal juvenile periods of 3.3–10.5 yr, which are in line with those we observed.

There was large variability in the degree of flowering of populations at given plant ages among sites and within species. In B. hookeriana (in western Australia), within and between year variability in cone production increased with plant age (Enright et al. 1996). In our study, the number of flowerheads produced per flowering plant did not vary widely within or among years in Protea species, although it was more variable in the single year that L. eucalyptifolium flowering was recorded. Results from the two sites where we undertook recurrent surveys suggest that flowering status is more closely related to plant height than to plant age (cf. Le Maitre & Midgley 1991). The substantial difference between the two sites in the plant age to maturation is unlikely to be due to species differences, given the lack of a species effect observed at the moist site as well as in our one-off surveys of flowering status. Site differences thus appear to play a key role in plant growth and maturation rates. The most noticeable abiotic difference between the two sites was aspect, with plants on the relatively dry, western slope being slower to grow and mature than those on the relatively moist, southern slope. Evidence for the effects of aspect (dry, northern or western slopes vs moist, southern or eastern slopes) on flowering status as measured in one-off surveys was, however, ambiguous.

Post-fire recruitment

Recruitment was always above replacement levels following fires that burned in vegetation of ≥7 yr post-fire age. The lack of a relationship between recruitment success and pre-fire vegetation age suggests that once a critical post-fire age (and by implication, seed bank size) is attained, factors other than seed abundance affect recruitment success. These factors may include season of fire, slope and aspect in relation to moisture regimes, pre-fire parent density and interspecific competition (Bond et al. 1984, 1995; Le Maitre 1988a,b; Mustart & Cowling 1993; Laurie & Cowling 1994; Heelemann et al. 2008, 2010).

There was considerable variation in recruitment success for any given FRI, species or site (cf. Bond et al. 1984, 1995; Midgley 1989; Laurie & Cowling 1994). Average to very good (including the highest of all records) recruitment occurred at the site with the longest FRI (38 yr). This finding is at variance with that of Bond (1980) in the Swartberg Mountains (ca. 100 km inland from our study area), where senescence (40–45 yr of age) negatively affected post-fire recruitment of fynbos. We concur with van Wilgen et al. (2011) that the occurrence of very old vegetation is not a key concern in the ecological management of fire in montane proteoid fynbos, both because it is very limited in extent (Bond 1980; T. Kraaij, unpublished data), and because recruitment does not appear to be negatively affected by relatively long inter-fire periods.

Minimum fire return interval

Our results on post-fire recruitment success of proteoids suggest that FRIs of <6 yr would result in the local extirpation of this guild from eastern coastal fynbos, whereas recruitment above replacement levels consistently should occur after fires at ≥7-yr intervals. Application of Kruger & Lamb's (1978) rule of thumb (that 50% of the individuals in a population should have flowered for at least three seasons) to observed proteoid flowering status (50% of plants flowered once by 5–6 yr of age; Fig. 2a), and assuming that plants flower every year after first flowering and that seeds take 7 mo to ripen (van Staden 1978), implies a minimum FRI of 9 yr for Tsitsikamma proteoid fynbos. We do not have data on the flowering status of populations at a post-fire age of 9 yr in order to support this estimate, but most, although not all, species surveyed had reached the required level of flowering at 11 yr post-fire (Fig. 2c; Table 1). Using twice the primary juvenile period (in our case, twice >4 yr; Figs 2a and 3) as a guide for minimum FRIs (Bell 2001), suggests a lower threshold in excess of 8 yr. Substantial variation, both in flowering status and post-fire recruitment, as well as disparity among estimates of minimum FRI based on these measures, emphasize the need to empirically verify rules of thumb that are currently used as fire management guidelines. Verification may be done by relating pre-fire flowering status of proteoid populations to their post-fire recruitment response at corresponding sites (as done for the site that burned at 5 yr of age).

All sites from which we collected data to estimate juvenile periods were located in the Tsitsikamma region. Juvenile periods (and by implication, minimum FRIs) are likely to be considerably longer in dry habitats where plant growth rates are lower (Le Maitre & Midgley 1992). This is substantiated by the difference observed in growth rates and flowering status between the sites on a relatively dry western and the moist southern slope, respectively. At a regional scale, rainfall is lower in the Outeniqua than in the Tsitsikamma region, and at Outeniqua Nature Reserve (immediately west of the study area) the time needed for 50% of individuals to flower once and three times was longer than that in the Tsitsikamma region (Table 1). Median FRIs, recorded since 1980, have accordingly been shorter in the Tsitsikamma than in the Outeniqua region (T. Kraaij, unpublished data). Overall, maturation rates of proteoids are comparable between the eastern and western CFK (Table 1), suggesting that variation is related to local moisture regimes (Kruger & Bigalke 1984) rather than a general east–west gradient within the CFK. Seydack et al. (2007) accordingly found an inverse relationship between FRI in fynbos shrublands and rainfall (related to plant productivity) along an altitudinal gradient in the Swartberg Mountains.

Range-restricted species or habitat specialists may have very specific FRI requirements (Kruger & Bigalke 1984). For instance, P. grandiceps (Near Threatened; Raimondo et al. 2009), a slow-growing, high-altitude species from the study area and elsewhere in the CFK, is slow to mature (>10 yr; J.H.J. Vlok, pers. comm., 2012, local botanist) and favours rocky outcrops and steep slopes where it is relatively safe from frequent fires (Rebelo 2001). Some subpopulations of this species have been exterminated as a result of too frequent fires (J.H.J. Vlok, pers. comm., 2012, local botanist). A survey done on the Kammanassie Mountains (30-km inland of our Outeniqua region) showed that only ten out of 200 plants flowered for the first time at 9 yr post-fire (CapeNature unpublished data).

Bradstock & Kenny (2003) discuss the limitations of using juvenile periods and life spans of the most sensitive plant guilds to inform boundaries for FRIs, and inter alia state that consistent favouring of one guild may in the long term lead to a loss of biodiversity. However, retention (and even dominance) of proteoids was shown to be key to the conservation of diversity in fynbos overall (Vlok & Yeaton 1999). Experience furthermore suggests that rigid control over fire regimes is largely unattainable (Keeley et al. 1999; Moritz 2003; van Wilgen et al. 2010), with wildfires providing sufficient variation to preclude consistent favouring of a particular plant guild.

In the same way that plant characteristics can be used to establish thresholds for FRIs, the post-fire development of habitat and fuel attributes (Haslem et al. 2011), and the life cycles and behaviour of selected animal species, may be used (Gill & McCarthy 1998). In the mediterranean shrublands of SW Australia, population modelling based on the demography and behaviour of rare, poorly-dispersing, ground-dwelling bird species suggested optimum FRIs in excess of plant maturation rates (Brooker & Brooker 1994; Gill & McCarthy 1998). South African fynbos does not have many endemic bird species, but some may be adversely affected by short FRIs. Both the Cape Sugarbird (Promerops cafer) and Protea Seedeater (Crithagra leucopterus) require mature proteoid fynbos for feeding and breeding, and it may take ≥10 yr post-fire for fynbos to attract breeding birds (Milewski 1978; Martin & Mortimer 1991). The existence of adequate areas of mature fynbos in the landscape is thus a requisite for these birds' persistence. While the focus of this paper has been on establishing minimum FRIs based on plant attributes, it would also be important to ensure that at least a proportion of the vegetation is in the age classes of 10–20 yr to conserve these faunal elements. This implies that the minimum age for burning would have to be >10 yr for a proportion of the area.

Management implications

The suggested minimum FRI of 9 yr in relatively moist and productive Tsitsikamma fynbos is not intended to prescribe rigid management of fire according to a fixed burning rotation. Neither does it negate the need to consider site- or species-specific requirements. Instead, it provides a lower threshold for a range of acceptable FRIs beyond which a significant decline of species populations is predicted (Bradstock & Kenny 2003; van Wilgen et al. 2011). Where two or more species of slow-maturing reseeders co-exist, the FRI for maximizing survival or abundance may be different (Gill & McCarthy 1998). Variation in fire regimes is therefore necessary to maintain plant diversity in the landscape (Cowling 1987; Gill & McCarthy 1998; Thuiller et al. 2007). In light of climate change and the associated increases in fire frequency that have been recorded locally (Forsyth & Wilgen 2008; Wilson et al. 2010; T. Kraaij, unpublished data) and in temperate shrublands globally (Piñol et al. 1998; Williams et al. 2001; Keeley & Zedler 2009), managers attempting to maintain fire regimes within ecological thresholds should follow a precautionary approach, particularly at the lower end of the FRI range. In addition to allowing for variation in the fire regime, they should aim for mean FRIs towards the middle of the ecologically acceptable range, rather than at the lower end, as the increasing occurrence of unplanned fires is likely to reduce mean FRIs overall.

Both frequent and low-intensity fires in fynbos and other temperate shrublands favour resprouters over slow-maturing, serotinous or myrmechocorous reseeders, leading to a loss in diversity overall (Haidinger & Keeley 1993; Vlok & Yeaton 1999, 2000). In large parts of the GRCMs, graminoid sprouters dominate, while proteoids are sparse or absent (Heelemann et al. 2010), notably in areas near plantations of alien pine trees (T. Kraaij, pers. obs.). This likely resulted from frequent and low-intensity burning in the past, aimed at protecting timber plantations from fire (Kraaij et al. 2011). Facilitation of FRIs that would ensure the long-term persistence of proteoids in the GRNP is therefore a priority for fynbos conservation management, particularly where this guild has persisted in the landscape. Short interval fires, which alter vegetation structure (from woody to herbaceous; Kruger 1984; Lloret et al. 2003) and thus fuel dynamics, may set up negative feedback loops whereby short FRIs persist in the landscape (Haidinger & Keeley 1993; Milton 2004).

Fynbos vegetation in the GRCMs is currently severely threatened by widespread invasion by pine trees (Pinus pinaster and P. radiata) grown in plantations that are scattered throughout the landscape (Cowling et al. 2009; Kraaij et al. 2011). Like the proteoids, pine trees are serotinous and fire-adapted, and repeated fires drive their rapid spread and proliferation (Richardson 1998). Prolonging the FRI (i.e. reducing the fire frequency) would thus generally be in the interest of invasive plant control by curbing the rate of spread of pines. On the other hand, in areas where the proteoids have already been lost due to past management practices, application of a single FRI that is shorter than the juvenile period of these pines (ca. 5–6 yr; Richardson et al. 1990) may in some instances provide an inexpensive means of substantially reducing dense infestations of young pine recruits.

Acknowledgements

Jan Vlok made conceptual contributions to this study and commented on an earlier draft of the manuscript. Comments of two anonymous reviewers resulted in improvements to the manuscript. Dr W.J. Vermeulen initially set up the survey site at De Vasselot Nature Reserve. Johan Huisamen, Diba Rikhotso, Johan Baard and CapeNature rangers assisted with field surveys. South African National Parks, the Nelson Mandela Metropolitan University, the Council for Scientific and Industrial Research and the South African Environmental Observation Network funded this study.