• Australia;
  • climate change;
  • mangrove;
  • range expansion;
  • salt marsh;
  • South Africa;
  • South America;
  • temperature;
  • USA


  1. Top of page
  2. Abstract
  3. Introduction
  4. Materials and methods
  5. Results
  6. Discussion
  7. Acknowledgements
  8. References

Mangroves are species of halophytic intertidal trees and shrubs derived from tropical genera and are likely delimited in latitudinal range by varying sensitivity to cold. There is now sufficient evidence that mangrove species have proliferated at or near their poleward limits on at least five continents over the past half century, at the expense of salt marsh. Avicennia is the most cold-tolerant genus worldwide, and is the subject of most of the observed changes. Avicennia germinans has extended in range along the USA Atlantic coast and expanded into salt marsh as a consequence of lower frost frequency and intensity in the southern USA. The genus has also expanded into salt marsh at its southern limit in Peru, and on the Pacific coast of Mexico. Mangroves of several species have expanded in extent and replaced salt marsh where protected within mangrove reserves in Guangdong Province, China. In south-eastern Australia, the expansion of Avicennia marina into salt marshes is now well documented, and Rhizophora stylosa has extended its range southward, while showing strong population growth within estuaries along its southern limits in northern New South Wales. Avicennia marina has extended its range southwards in South Africa. The changes are consistent with the poleward extension of temperature thresholds coincident with sea-level rise, although the specific mechanism of range extension might be complicated by limitations on dispersal or other factors. The shift from salt marsh to mangrove dominance on subtropical and temperate shorelines has important implications for ecological structure, function, and global change adaptation.


  1. Top of page
  2. Abstract
  3. Introduction
  4. Materials and methods
  5. Results
  6. Discussion
  7. Acknowledgements
  8. References

The increase in global average surface temperature of 0.74 °C (1906–2005) (Solomon et al., 2007) has already caused shifts in the structure and distribution of ecological communities at a variety of scales (Walther et al., 2002; Parmesan & Yohe, 2003). Arctic shrubs have advanced northward in response to decreases in intensity of freezing (Sturm et al., 2001), and an advance in range has been demonstrated for butterfly species (up to 200 km) (Parmesan et al., 1999) as well as birds (an average of 20 km for 12 bird species in Britain) (Thomas & Lennon, 1999). Minimum temperatures globally are increasing at twice the rate of maximum temperatures (Walther et al., 2002). In temperate climates, increasing temperature and decreasing intensity and frequency of frost are likely to cause transitions in the distribution of temperature sensitive higher plants (Bakkenes et al., 2002; Loarie et al., 2008), which in many instances provide structural habitat and organic carbon to organisms and ecosystems.

In many ways, mangroves are ideal species for monitoring the impacts of global climate change on vegetated habitats. Mangroves are sensitive to several global environmental conditions undergoing change, including enhanced atmospheric CO2 (McKee & Rooth, 2008), sea level (Woodroffe, 1990; McKee et al., 2007), temperature (Alongi, 2008), and rainfall (Semeniuk, 2013). All mangrove species are hydrochorous and thus often have some potential for dispersal to new localities by sea currents and drift (see Friess et al., 2012; Van der Stocken et al., 2013). Mangroves are conspicuous and can be identified from aerial photography at a scale represented in easily accessible geographic applications such as Google Earth ( and Nearmap (, displaying an emergent canopy above salt marsh in temperate and subtropical intertidal environments, although on-ground verification may be required when grading to freshwater woody vegetation. They are an important habitat for estuarine, nearshore and terrestrial biota (Nagelkerken et al., 2008), and play a critical role in coastal environments in stabilising shorelines (Gedan et al., 2011), and sequestering atmospheric carbon (Chmura et al., 2003; Donato et al., 2011).

Temperature has long been considered the primary limit to the latitudinal range of mangroves. Walsh (1974) postulated that this poleward threshold corresponded to a mean monthly atmospheric temperature of 20 °C for the coldest month. Duke et al. (1998) more accurately identified the winter position of the 20 °C isotherm for sea surface temperature (SST) as corresponding to the latitudinal limit in both hemispheres (Fig. 1), although SST and air temperature at the latitudinal limit of individual species and genera may vary between continents (Quisthoudt et al., 2012). While mean temperatures provide a correlative explanation for mangrove distribution, quantifying minimum temperature requirements (and measures of extreme winter events) provide an even better mechanistic approach for quantifying thresholds (Osland et al., 2013). That mangroves will shift their distribution after meeting minimum temperature thresholds in response to changing climate is well attested by the fossil record. Mangrove species distribution has changed in concert with small changes in temperature since the early Holocene. For example, a slight cooling following the mid-Holocene highstand (6000 years bp) is associated with the less common occurrence of Rhizophoraceae in northern NSW (Hashimoto et al., 2006), and the loss of Avicennia marina from the Poverty Bay-East Cape region of New Zealand (Mildenhall, 1994).


Figure 1. Global mangrove and salt marsh distribution and the average 20 °C sea-surface temperature isotherm. Sources: Spalding (2012), Hoekstra et al. (2010), and NOAA (2013).

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However, caution should be exercised in interpreting changes in distribution and latitudinal limits solely to temperature. The effects of temperature upon mangroves are mediated by interactions with other aspects of global change (e.g., CO2, precipitation, sea-level rise, nutrients). Geomorphic changes in response to rising, and then stabilising sea level exerted the strongest control on mangrove extent over the Holocene (e.g., Grindrod et al., 1999; Hashimoto et al., 2006). Both fluctuating sea levels and temperature regimes have vastly influenced mangrove distributions globally since much older geological time frames than the Holocene (Sherrod & McMillan, 1985; Ellison et al., 1999). Contemporary distributions are shaped by suitable intertidal habitat, and the capacity of floating propagules to access these locations. Impediments to colonization therefore include unfavorable ocean currents, closed estuary entrances, or on arid and hard-rock coastlines, an absence of estuaries with depositional environments suitable for mangrove establishment (Saintilan et al., 2009). Such impediments have slowed the filling of potential niche as defined by temperature thresholds for many species (Quisthoudt et al., 2012).

Several publications have postulated that mangroves will migrate to higher latitudes, replacing salt marsh as an outcome of global warming (Woodroffe & Grindrod, 1991; Field, 1995; Gilman et al., 2008). However, assessments of changes in mangrove extent at poleward limits are restricted to a few site specific studies. In this paper, we use published historic records of occurrence and distribution limits, contemporary published surveys, and our own observations to provide a global synthesis of evidence for proliferation and extension of mangroves at poleward limits. Mangroves are absent from Europe and the Mediterranean Sea but coexist with salt marsh in temperate settings in Asia, Africa, Australia/New Zealand, North America, and South America.

Materials and methods

  1. Top of page
  2. Abstract
  3. Introduction
  4. Materials and methods
  5. Results
  6. Discussion
  7. Acknowledgements
  8. References

We present a synthesis of literature from four continents detailing changes in the distribution of mangroves. In some cases, we provide additional analyses using remote sensing, field survey and local expert observations. We used Google Earth Pro ( to confirm occurrence within estuaries and poleward extent in each of the focus regions using the most recent available imagery. These images included photography of Cedar Keys, Florida (imagery dated 19 January 2012), Virilla estuary, Peru (imagery dated 19 January 2010, DigitalGlobe), and Piura estuary, Peru (imagery dated 10 February 2011, DigitalGlobe). We also used Google Earth Pro's polygon area function to estimate the extent of mangroves where these had expanded from the time of previously published estimates, including an update of the estimates in Stevens et al. (2006) for the US Gulf Coast, and the area of mangroves in Piura, Peru. We interpreted mangrove and salt marsh using techniques defined in Wilton & Saintilan (2000). Our identification of mangroves in Vichayal, Peru using Google Earth Pro was confirmed by photographs provided by Edwin Gerardo and Manuel Ravelo.


  1. Top of page
  2. Abstract
  3. Introduction
  4. Materials and methods
  5. Results
  6. Discussion
  7. Acknowledgements
  8. References

Northern hemisphere

North America

Mangroves occupied intertidal locations in the southeastern USA at least as far back as the early Eocene Epoch (ca. 45 Million years bp), but those fossil deposits were associated with a vastly different coastline boundary driven by a warmer climate and higher sea level (Berry, 1916, 1924; Westgate & Gee, 1990). Mangrove forests from the Eocene Epoch likely occurred at densities similar to those seen in modern-day Neotropical mangrove forests, just much farther north (Sherrod & McMillan, 1985; Gee, 2001). The first fossil evidence of Avicennia in the Caribbean appeared in the late Miocene Epoch (ca. 10 Million years bp), and by the mid-Pliocene Epoch (ca. 3.5 Million years bp) multiple mangrove genera were evident (Graham, 1995). A prominent lack of mangrove fossil evidence along the northern Gulf just preceding the Pleistocene Epoch (ca. 11 700 years bp) until 3000–4000 years bp (from Holocene peat deposits in south Florida) suggests an eradication event for mangroves along the northern Gulf of Mexico, perhaps related to colder temperatures when mangroves were aligned in distribution closer to the equator (Sherrod & McMillan, 1985).

At the northern limits of present-day mangrove extent in the Gulf of Mexico, population extent has, in the recent past, been periodically reduced by frost (McMillan & Sherrod, 1986), with heavy frost in 1983 and 1989 leading to 95–98% loss amongst several of the northernmost populations (Lonard & Judd, 1991; Everitt et al., 1996; Montague & Odum, 1997). This observation prompted Snedaker (1995) to suggest that periodic heavy frost would limit northern expansion for some time. Ecotypic differences in cold tolerance among natural mangrove populations in the Gulf do have the potential to buffer this impact somewhat. This is especially true for populations of Avicennia germinans (McMillan, 1971); those populations growing along the Texas coast were especially tolerant to freezing among others surveyed in the wider Caribbean region (Markley et al., 1982). However, in more than 20 years since the 1989 freeze event, winters have been sufficiently mild to allow rapid expansion of mangroves at their northern limits into salt marsh, documented in Texas (Comeaux et al., 2012; Bianchi et al., 2013), Louisiana (Perry & Mendelssohn, 2009; Alleman & Hester, 2011; Pickens & Hester, 2011) and Florida (Stevens et al., 2006).

Avicennia germinans coverage increased from 57 ha in 1986 to 1182 ha in 2006 in Louisiana, but fluctuated from a maximum documented coverage of approximately 2180 ha in 1983 before the freeze of that same year (Giri et al., 2011). By another account, A. germinans increased in abundance by nearly fivefold between 2002 and 2009 within the Louisiana deltaic plain (Michot et al., 2010). Populations of A. germinans seem to be regulated strongly by air temperatures of −6.7 to −8.9 °C or less (Lonard & Judd, 1991; Stevens et al., 2006; Osland et al., 2013). This threshold is more restrictive for other Neotropical mangrove species (Lugo and Patterson Zucca, 1977; Krauss et al., 2008). For instance, there was no reported survival of transplanted Rhizophora mangle seedlings after the 1983 freeze in Texas (Sherrod et al., 1986), and embolism is a common consequence of temperatures slightly below 0 °C in the same species (Fig. 2a and b). Likewise, Laguncularia racemosa trees are highly susceptible to repetitive freeze-induced dieback events (Fig. 2c), although re-sprouting from the base is a common response in both L. racemosa and A. germinans.


Figure 2. (a) Air temperatures (°C) for the Ten Thousand Islands region of Florida, USA from November 2006 through April 2007, with days having subzero temperatures highlighted (inset graphs). These subzero temperatures were responsible for (b) branch tip mortality from vascular embolism in Rhizophora mangle, and (c) complete stem dieback in many Laguncularia racemosa trees growing in open environments. Avicennia germinans trees in the Ten Thousand Islands region were generally unaffected by this freeze. (Temperature data source: DBHYDRO Browser, South Florida Water Management District,, Station SGGEWX, accessed 11 April 2013).

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Indeed, after extensive losses during the 1983 freeze, mangroves have extended in many Gulf study sites since 1984 (Giri et al., 2011) although have not reached pre-1983 extent (C. Giri, unpublished results). Mangrove trees have been documented visually in Louisiana as early as 1938 (Penfound & Hathaway, 1938) and in Texas as early as 1853 (cited in Sherrod & McMillan, 1981). Using an historical time-series of aerial photography extending back to 1956, Perry & Mendelssohn (2009) were able to demonstrate that mangroves first occupied their Louisiana site in 1995. Along with a reduced incidence of freeze-induced mortality, recent expansion of mangroves in Louisiana has been assisted by widespread dieback of S. alterniflora resulting from drought; Avicennia germinans was unaffected by drought and proliferated (McKee et al., 2004). Environmentally mediated competition between S. alterniflora and mangroves also occurs along latitudinal gradients in Florida (Kangas & Lugo, 1990) and was probably of importance during post-Pleistocene recolonization of mangroves toward northern latitudes. A recent analysis applied to the northern Gulf suggests that short-statured A. germinans vegetation has an overall lower requirement for water use in early growing season assessments than S. alterniflora (Krauss et al., 2013). This may help to explain the differential survival of A. germinans over S. alterniflora during drought, and suggests an interaction between climate variability in both temperature and rainfall (Krauss et al., 2013).

Much of what we are now documenting in the Southern USA is the northern boundary of the post-Pleistocene recolonization (sensu Sherrod & McMillan, 1985). Currently, mangroves (primarily A. germinans) have also extended north on the Florida Atlantic coast at least as far as St Augustine, occupying back-barrier intertidal flats as scattered clusters of individuals (29°57′59″N), and have expanded within this estuary since the early 1990s. In fact, A. germinans has expanded into salt marsh at several other sites on the Atlantic coast, including the Indian River lagoon (Harris & Cropper, 1992). To the south, Rhizophora mangle has expanded landward more than a kilometer into previously Cladium and Eleocharis marshlands in the Everglades (Ross et al., 2000), possibly in response to higher sea levels, changing water levels, and shifting fire regimes (Smith et al., 2013). Similar landward expansion has been noted on the Pacific coast of Mexico at Magdalena Bay, Baja California. Here, a 20% increase in mangrove extent through landward encroachment into sparse halophytic shrubland was attributed to sea-level rise, and was particularly pronounced during El Nino seasons (Lopez-Medellin et al., 2011).

On the Gulf Coast of Florida, mangroves increased coverage in the Ten Thousand Islands National Wildlife Refuge by 35% since 1927, principally at the expense of salt marsh (Krauss et al., 2011). Over a similar time period, oscillations between marsh and mangrove area have been documented in other Gulf coastal areas of Florida (Egler, 1952; Bischof, 1995; Smith et al., 2013); sometimes to the detriment of marsh and sometimes to the detriment of mangroves. In the absence of any discernable change in mean number of freeze days over the period, encroachment of mangroves onto marsh was attributed primarily to the increase in sea level over the period (2.24 mm yr−1 at the Key West station: Krauss et al., 2011). After comparing mangrove extent at three sites in Cedar Keys between 1995 and 1999, Stevens et al. (2006) predicted that all three sites would develop complete mangrove cover within 25–30 years, if not impacted by frost. Our assessment of the same sites using 2012 aerial photography (Google Earth imagery, 19 January 2012) suggests that this outcome has been realized in less than half the predicted time.


There are insufficient historic data on the southeast Japanese coast to unequivocally argue for an extension in natural range of Kandelia obovata (syn. K. candel). The northern limit of K. obovata in Japan was reported by Wakushima et al. (1994) to be Kiire, Kagoshima Prefecture (31°30′N), although they note the long-term survival of a planted population in the estuary of the Aono River in the Shizuioka Prefecture at 34°38′N.

Determining changes in northern limits of mangroves in China and Taiwan is complicated by extensive clearance. A further complication in China is the introduction of mangroves north of their natural limits: K. obovata in Zheihang (Li & Lee, 1997); and Sonneratia caseolaris and Bruguiera sexangula in Guangdong (Li et al., 1998). One of the few locations where mangroves and salt marshes coexist in near natural state on the Chinese mainland coast is in the Zhanjiang Mangrove National Nature Reserve on the Leizhou Peninsula of Guangdong Province (21°34′N; 109°45′E). The reserve is a Ramsar-listed wetland of international significance and supports nearly one-third of China's mangroves. Regionally, mangroves have declined due to agricultural developments, and extensive dyking restricts landward encroachment (Leempoel et al., 2013). However, within the reserve mangroves, dominated by A. marina, Aegiceras corniculatum and K. obovata, have expanded fourfold, including encroachment on salt marsh (Durango-Cordero et al., 2013). Mangroves have also proliferated in the Zhuhai Qi'ao Provincial Nature Reserve (22°26′N; 113°37′E), established in 2000 to encourage the rehabilitation of mangroves (Peng et al., 2009). Spread in the extent of the native mangrove K. obovata as well as Sonneratia apelata, introduced from the Sunderban (Ren et al., 2009), has led to a decline in Spartina alterniflora saltmarsh (G. Lei, personal communication).

The northernmost mangrove community in Taiwan is located in the Danshui River estuary (21°09′N; 121°26′E) and is the largest K. obovata forest in the world (Lee & Yeh, 2009). The mangrove and associated Phragmites communis salt marsh community has been protected in the Danshui Mangrove Reserve since in mid-1980s. Mangroves have doubled in extent since the establishment of the reserve, and in detailed satellite imagery analysis Lee & Yeh (2009) were able to demonstrate landward encroachment of mangrove on nonmangrove vegetation, presumably Phragmites salt marsh.

Southern hemisphere


The gray mangrove A. marina extends south on the Australian mainland to the southernmost intertidal flats within Corner Inlet, Victoria (38°54′25″S), and has occupied this range since the earliest historic records from the 19th century. These are the southernmost mangroves in the world, and the Bass Strait provides an effective barrier to further dispersal to the north coast of Tasmania. A. marina in southern Australia is exposed to more frequent but less extreme frosts than those encountered in the US Gulf Coast by A. germinans, and has developed a greater resistance to freeze-induced embolism (Stuart et al., 2007).

Mangrove expansion within estuaries is a near ubiquitous trend in southeastern Australia, (Saintilan & Williams, 1999), and New Zealand (Burns & Ogden, 1985; Morrisey et al., 2003; Lovelock et al., 2007; Stokes et al., 2010), and has been occurring since the time of earliest aerial photographic records (1950s), and perhaps earlier (McLoughlin, 1988, 2000). Temperature increases across the region over the past century are likely to be one of a suite of regional environmental changes promoting mangrove growth and a corresponding loss of salt marsh, including sea-level rise (Rogers et al., 2006), increases in sedimentation following catchment development (emphasized in New Zealand studies: Lovelock et al., 2007; Swales et al., 2007; Morrisey et al., 2010) and, in Queensland, higher rainfall (Eslami-Andargoli et al., 2009). Mangroves in New Zealand have expanded across 29 locations by an average of 165% since the 1940s. There is less obvious salt marsh decline than in Australia (Morrisey et al., 2010), possibly due to higher sedimentation rates and elevation gain (Stokes et al., 2010), although some landward encroachment has been noted (Burns & Ogden, 1985). A median estimate of 30% of salt marsh has been lost to mangrove encroachment across SE Australia (Saintilan and Williams 2000; Straw & Saintilan, 2006), with some evidence that rates of loss are lower toward the southern limit in Victoria (5–10%) (Rogers et al., 2005), although this may be due to competitive resilience of large saltbushes of the genus Tecticornia, as much as colder conditions slowing mangrove expansion in the south.

Mangrove floristic diversity declines with increasing latitude on the east and west coasts of the Australian continent. On the west coast, patterns in mangrove diversity at a regional scale are strongly influenced by aridity, confounding the assessment of temperature effects on mangrove species range expansion (Semeniuk, 1983; Wells, 1983). The humid subtropical-temperate east coast presents an ideal setting to explore changes in mangrove diversity, with a cline in temperature extending across more than 150 estuaries, linked by the south-flowing East Australia Current south of the Great Barrier Reef. Species of the tropical family Rhizophoraceae (Rhizophora stylosa and Bruguiera gymnorrhiza) were common in northern NSW during the early- to mid-Holocene, when temperatures and sea levels were likely to have been higher than present (Hashimoto et al., 2006), although were rare in the earliest contemporary surveys (Wells, 1983; West et al., 1985) with R. stylosa recorded in seven estuaries in northern NSW. Both R. stylosa and B. gymnorrhiza appear to have expanded their range in recent decades. Bruguiera gymnorrhiza has recently colonized at least three southerly estuaries, the Sandon, Wooli Wooli Rivers, and Moonee Creek (Wilson, 2009). Rhizophora stylosa has now been recorded within 16 estuaries (Wilson, 2009), and has shown strong population growth within a number of NSW estuaries (Wilson & Saintilan, 2012). Although it is highly probable that R. stylosa was missed in at least two estuaries in earlier surveys in NSW, the colonization of others is clearly very recent, based on demographics. The 100 km southward extension of R. stylosa from the Corindi estuary to South West Rocks Creek (30°53′16″S), corresponds to the southward shift in temperature zones in the region over the past few decades (Hennessy et al., 2004). However, colonization of estuaries between these latitudes is sporadic rather than incremental, and leaf phenology does not suggest a temperature cline limiting growth (Wilson & Saintilan, 2012).

South Africa

The earliest comprehensive survey of South African mangroves now dates back 50 years, and represents aerial photographic and field surveys over a 14-year period to 1962 (Macnae, 1963). South of Port St Johns, Macnae (1963) reported stands of mangroves at the estuaries of the Mtata (29°11′E, 31°57′S) and Mngazana Rivers (29°25′E, 31°42′S), ‘isolated clumps’ of mangroves at the estuaries of the Mbashe (29°25′E, 31°42′S) and Nxaxo (28°31′E, 32°35′S) Rivers, and ‘occasional trees’ southward. Macnae (1963) reported temperature thresholds on the basis of his observations of distribution as being 19 °C mean air temperature or where the mean of the coldest monthly air temperature does not drop below 13 °C. This placed the Mbashe and Nxaxo estuaries at the southern limit (19.1 °C mean, 11.9 °C mean coldest monthly), with Bufallo River in East London outside of the range (17.7 °C mean, 10.2 C mean coldest monthly).

Mean temperature at the Buffalo River for the period 1973–2011 rose from 17.7 to 18.7 °C, and the mean coldest temperature rose from 10.2 to 14.4 °C (Tutiemp, 2012), a shift extending the possible range of mangrove in South Africa to East London based on the untested thresholds of Macnae (1963). Some dispersal challenges on the Transkei coast include the proportionately high number of temporarily open/closed estuaries (17 of the 76 estuaries are permanently open), and although the Agulhas current flows south 2–3 km offshore, a counter-current develops between the Agulhas and the shoreline creating a predominantly northward drift (Macnae, 1963). In spite of these challenges, and widespread clearing of mangroves, in the 20 years to 1982 mangroves formed extensive stands in the estuaries of the Kobonqaba (28°30′E, 32°36′S, to the south of the Nxaxo), Nqabara (28°47′E, 32°30′S), Xora (29°05′E, 32°05′S), and Bulungula (29°00′E, 32°08′S) Rivers (Ward & Steinke, 1982). It is unlikely these were missed by Macnae; mangroves cover a larger area on the Xora estuary (16 ha) than the Mbashe (12.5 ha) and Nxaxo (14 ha), and line the lower shore of the estuary. In 1969, mangroves were observed for the first time in the Kwelera River (32°54′S, 28°04′E), still the southernmost known natural stand. Natural seeding in the Kwelera River is strongly suggested by the results of a drift card dispersal experiment, in which one of the cards dropped offshore of the Nxaxo River was retrieved within 100 m of the Kwelera mangrove stand (Steinke & Ward, 2003).

Mangrove area has increased by approximately 40% in South African estuaries since the 1970s, with most of the gains in the Umhlatuze estuary (increase from 197 to 489 ha: Bedin, 2001; Ward & Steinke, 1982) and the Mtata (increase from 34 to 42 ha 1982–1999: Adams et al., 2004). Small declines were observed in more than half of estuaries sampled by Adams et al. (2004), and mangroves have been lost entirely from many estuaries (Quisthoudt et al., 2013). This may be related to limited available habitat for colonization (Wright et al., 1997) and in some cases the removal of mangroves manually (the Mnyameni: Adams et al., 2004) but is principally attributed to prolonged inundation following long-term closure of the estuary mouths on temporarily open/closed estuaries (e.g., the Bulungula, Umzimvuba, Kosi and Kobonqaba rivers: Breen & Hill, 1966; Adams et al., 2004).

However, mangroves appear to have established naturally in the Kei River (28°21′42″E, 32°40′00″S,) to the north of the Kwelera, and the Gqunube River (28°02′E, 32°56′S) to the south, with the Kobonqaba River a possible source (Steinke, 1986; Steinke & Ward, 2003). It is uncertain whether the Gqunube River mangroves were naturally dispersed or planted.

Avicennia marina, B. gymnorrhiza, and Rhizophora mucronata have also survived in the Nahoon estuary in East London after being transplanted from Durban Bay (Steinke, 1999), suggesting that climate was or is no longer a factor limiting their southern natural extent. Of these three species, it is only A. marina that has expanded substantially within the estuary, and now covers 1.6 ha of previously salt marsh flat, and is expanding at 0.1 ha yr−1 (A. Rajkaran, personal observations 2012). Quisthoudt et al. (2013) were able to successfully predict current distribution of A. marina, B. gymnorrhiza, and R. mucronata based on current climate variables, with number of growing days above an 18 °C threshold being the most important. On this basis, they predict latitudinal expansion of mangroves with continued climatic warming.

South America

Mangroves grow south on the Atlantic coast to Santo Antonia Lagoon in the Municipality of Laguna (28°28′S; 48°50′W) (Soares et al., 2012). This southern limit has not changed in the two decades since the survey of Schaeffer-Novelli et al. (1990), although populations of the dominant species Laguncularia recemosa show evidence of recent recruitment (Soares et al., 2012). At this site L. racemosa is stunted, a trait in common with species globally at their southern limit, although Avicennia schaueriana grows to 10 m, suggesting a vigour characteristic of a species well within its range (Soares et al., 2012). Further southward expansion may be limited by a strong northerly current described by Siegle & Asp (2007) extending from Ararangua, an estuary 100 km south, to Laguna (Soares et al., 2012).

The southern limit of mangrove communities on the South American west coast was considered by Clüsener & Breckle (1987) to be the River Thumbes at 3°35′S; beyond which were found only a few small individuals of Rhizophora near the village of Bocapan (at 3°44′S), and a small stand of Avicennia at the mouth of the Piura River. Mangroves were successfully planted within this range in their experimental studies in 1984–1985.

South of Cerro Illescas (6°0′S), the cold Peruvian current precludes mangrove colonization (Clüsener & Breckle, 1987), and because of the aridity of the coast, only three estuaries between Cerro Illescas and Bocapan provide intertidal conditions suitable for the development of mangrove, these being the Virrila estuary (5°50′S); the Piura River (5°30′S) and the Vichayal estuary (4°53′S). The ‘small stand’ of Avicennia described by Clüsener & Breckle (1987) at Piura is now very extensive, lining 9.5 km of shoreline and covering at least 38 ha in the north arm and 9 ha in the south arm of the estuary at San Pedro, the southernmost confirmed mangroves on the west coast (imagery dated 10 February 2011, DigitalGlobe, sourced from Google Earth Pro). The Vichayal estuary has a new stand of Avicennia at 4°53′22.6″S; 81°08′56.4″W covering 1.87 ha (field photographs provided by Manuel Ravelo, imagery dated 19 January 2010, DigitalGlobe, sourced from Google Earth Pro). These are absent from aerial photographs taken in 1970 (Google Earth Pro) and reportedly established during the El Niño event in the first decade of this century (E. Gerardo, personal communication 2012).


  1. Top of page
  2. Abstract
  3. Introduction
  4. Materials and methods
  5. Results
  6. Discussion
  7. Acknowledgements
  8. References

Dispersal may be problematic in spite of the abundance of buoyant propagules produced by Avicennia spp. (Clarke et al., 2001; Sousa et al., 2007), and restricted gene flow in marginal populations (Dodd & Afzal Rafii, 2002) also suggests dispersal may restrict the expansion of range. In many places, the latitudinal limit of mangroves appears to lag behind changes in temperature thresholds, as documented in New Zealand (de Lange & de Lange, 1994), east coast Australia (Wilson & Saintilan, 2012), South Africa (Steinke, 1999), and South America (Soares et al., 2012). The difference between fundamental and realized niche is relatively large for Avicennia and Rhizophora on the basis of global comparisons (Quisthoudt et al., 2012), and on some coastlines may reflect slow expansion from Pleistocene extents. Disequilibrium between tree species distribution and rapidly changing temperature regimes has been noted for terrestrial species also (Willner et al., 2009). It is likely that a more complex response than a steady stepping poleward will be the case for many mangrove species, especially those on relatively high wave energy coasts with few permanently open estuaries or where dispersal is subject to unfavourable currents. This infers that there is no simple function relating range extension and warming temperatures, something also implied by the global temperature and range analysis of Quisthoudt et al. (2012).

Parmesan & Yohe (2003) found poleward range shifts in 75–81% of 1045 species of higher plants and animals with quantitative records, with an average shift of 6.1 km per decade. Notwithstanding limited opportunities for dispersal and the difficulties of ‘threading the needle’ of estuarine entrances, an increase in range has been documented for the mangroves A. germinans in the USA and Peru, A. marina in South Africa and R. stylosa and B. gymnorrhiza in eastern Australia; and expanding mangrove populations near poleward limits are obvious within estuaries in Australia, New Zealand, the Gulf and Atlantic coasts of the USA, the Pacific and South Atlantic coasts of South America, and the Leizhou Peninsula of China, one of the few locations in southern China where large areas of mangrove and salt marsh are protected and have been retained. Poleward expansion in the coming decades will be most evident on open coasts where temperature currently exerts a strong control on contemporary distributions and available habitat exists. Osland et al. (2013) used contemporary mangrove forest distribution data and 30-year climate records from the Gulf and Atlantic US coasts to identify winter-climate based thresholds and develop mangrove species distribution and relative abundance models. Their models and analyses of the potential effect of alternative future winter climate scenarios show that, in southeastern USA and especially in Louisiana, Texas, and Florida, relatively small changes in winter climate can result in relatively dramatic mangrove range expansion at the expense of salt marsh. Applying a 2–4 °C increase in annual mean minimum temperature would lead to a 95% reduction in salt marsh in Louisiana, 100% reduction in Texas and 60% reduction in Florida (Osland et al., 2013).

The comprehensive replacement of salt marsh by mangrove (cf., Osland et al., 2013; Guo et al., 2013) is predicated on temperature as the key delimiting factor of mangrove range expansion. In addition to temperature, local patterns of mangrove expansion into salt marsh are likely to be influenced by interactions between hydroperiod, sedimentation, elevation and salinity, with nutrients playing a role in some settings (Patterson & Mendelssohn, 1991; Patterson et al., 1997), all of which can be impacted locally by human agency, such as building walls and structures in estuaries, dredging, and development in the catchment. In coastal Louisiana, mangroves currently tend to dominate higher elevation settings such as the shorelines of tidal creeks, and exclusion from lower interior marshes has been attributed to higher predation, lower retention of propagules (Patterson et al., 1997), plant competition, and greater flooding stress (Patterson et al., 1993). By contrast, mangroves in eastern Australia show greater mortality in less frequently inundated higher salinity areas where propagules become desiccated (Clarke & Allaway, 1993; Clarke & Myerscough, 1993). That mangroves are invading salt marshes in contrasting settings along the northern Gulf of Mexico vs. Australia would suggest that different mechanisms are at work, or that global changes are contributing to an increased capacity of mangroves to survive in previously marginal intertidal environments.

Mangrove expansion into salt marsh mirrors a global trend of woody shrub invasion of grassland (Knapp et al., 2008; Williamson et al., 2010), which has been attributed variously to altered fire and grazing intensity (Scholes & Archer, 1997; Van Auken, 2009), and elevated atmospheric CO2 (Polley et al., 1997; Eamus & Palmer, 2008). On most coastlines, there is little evidence that altered fire and grazing regimes are dominant drivers of vegetation change in intertidal settings. The proliferation of mangroves in previously salt marsh-dominated environments is likely to be driven by a suite of environmental factors favoring mangrove and which are changing globally, including elevated sea level, elevated atmospheric CO2, and higher temperatures (Williamson et al., 2010; McKee et al., 2012). Landward encroachment of mangrove into salt marsh and salt pan has been attributed to sea-level rise in environments as disparate as Baja California (Lopez-Medellin et al., 2011), the US Gulf Coast (Krauss et al., 2011; Smith et al., 2013), and east coast Australia, where Rogers et al. (2006) demonstrated a lower capacity of salt marsh to respond to sea-level rise through vertical accretion. Salt marsh floristic diversity increases in inverse correlation with mangrove diversity on the Australian east coast (Saintilan, 2009) and mangrove encroachment may place further pressure on an ecological community already listed as endangered in New South Wales.

The replacement of salt marsh by mangrove in temperate settings has important implications for ecosystem organization and function. Experimental studies in the Gulf of Mexico (Comeaux et al., 2012) and temperate Australia (Rogers et al., 2006) show improved mineral trapping leading to a higher rate of surface elevation gain in encroaching mangrove than surrounding salt marsh, suggesting mangrove has greater potential to respond to increasing sea levels, although some of these differences may relate to different topographic settings. Carbon sequestration may be enhanced in some settings as a result of mangrove encroachment (Howe et al., 2009; Bianchi et al., 2013) and reduced in others, if redox potential is enhanced by mangrove root formation (Comeaux et al., 2012). The conversion of salt marsh to mangrove in the Gulf of Mexico alone could sequester 129 ± 45 Tg C over 100 years (Bianchi et al., 2013), more than 1% of ‘Blue Carbon’ estimates globally (Bianchi et al., 2013; Hopkinson et al. 2012), and a proportion that may rise if the trend of tropical mangrove deforestation continues (Valiela et al., 2001).


  1. Top of page
  2. Abstract
  3. Introduction
  4. Materials and methods
  5. Results
  6. Discussion
  7. Acknowledgements
  8. References

The authors thank Edwin Gerardo, and Dr J. Manuel Charcape Ravelo, National University of Piura, for their comments and photographs of mangroves in Peru. Robert J. Williams is thanked for comments on an earlier version of this manuscript. The use of trade, product, or firm names is for descriptive purposes only and does not imply endorsement by the US Government. We thank the UNEP World Conservation Monitoring Centre (WCMC) for access to mangrove and salt marsh distribution data, and Professor Guangchun Lei of Beijing Forestry University for his observations and photographs of mangrove encroachment into salt marsh in Guangdong Province, China.


  1. Top of page
  2. Abstract
  3. Introduction
  4. Materials and methods
  5. Results
  6. Discussion
  7. Acknowledgements
  8. References
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