2.1 Study area background: coastal Louisiana

Coastal Louisiana is positioned within the dynamic Mississippi River Delta, which is one of the largest river deltas in the world. In addition to providing valuable fish and wildlife habitat, Louisiana's abundant wetlands offer protection from storms, improve water quality, sequester carbon from the atmosphere, support food webs, provide seafood, store floodwaters and provide recreational opportunities (Barbier et al., 2011; Costanza et al., 2014). Though difficult to quantify, the societal benefits (i.e., ecosystem services) provided by coastal ecosystems in Louisiana's Mississippi River Delta have been valued to be at least $12–$47 billion (US dollars) per year (Batker et al., 2010). However, the rate of wetland loss in Louisiana has been very high in the past century, due to a combination of natural and human factors that have reduced the ability of wetlands to build elevation to keep pace with high rates of subsidence and relative sea-level rise (Blum & Roberts, 2009; Day et al., 2007; Törnqvist, Jankowski, Li, & González, 2020). Between 1932 and 2016, Louisiana lost approximately 4,833 km² of wetlands (Couvillion, Beck, Schoolmaster, & Fischer, 2017), and the state has become a prominent global example of the negative linkages between high relative sea-level rise, sediment delivery alterations and coastal wetland loss (Jankowski, Törnqvist, & Fernandes, 2017; Twilley et al., 2016). Given the rapid pace of wetland loss in the last century and the expectation of accelerated sea-level rise in the coming century, coastal scientists in Louisiana are increasingly challenged to better anticipate and prepare for the effects of climate change on Louisiana's coastal wetlands (Coastal Protection & Restoration Authority of Louisiana, 2017). However, Louisiana's coastal wetlands are vulnerable to interactions between many different aspects of climate change. For example, accelerated sea-level rise, intensifying hurricanes, more extreme drought, more extreme precipitation and winter warming are all aspects of climate change that are projected to play increasingly important roles in Louisiana (Carter et al., 2018; USGCRP, 2017, 2018).

Winter warming is a critical component of climate change that is expected to transform Louisiana's coastal ecosystems (Gabler et al., 2017; Osland, Day, et al., 2017; Osland, Enwright, Day, & Doyle, 2013). Extreme freeze events have historically served as major ecological disturbances in Louisiana and other tropical–temperate transition zones across North America—leading to mass mortality of freeze-sensitive organisms. Freeze events control the distribution of mangrove forests in Louisiana (Osland, Day, et al., 2017) and across the northern Gulf of Mexico and Atlantic coasts of North America (Cavanaugh et al., 2014; Osland et al., 2013; Ross et al., 2009; Sherrod & McMillan, 1985; Stevens, Fox, & Montague, 2006). Most of Louisiana's salt-affected tidal wetlands (i.e., those in the saline, brackish and intermediate salinity classes) are currently dominated by freeze-tolerant salt marsh graminoid plants (i.e., grasses, sedges and rushes; Osland, Grace, et al., 2019; Sasser, Visser, Mouton, Linscombe, & Hartley, 2014; Visser, Sasser, Chabreck, & Linscombe, 1998, 2000). Due to their current and historical role as critical foundation plant species (sensu Ellison, 2019; Ellison et al., 2005), these graminoid-dominated salt marsh plant communities tend to be the target of future-focused ecological assessments and restoration-focused planning efforts for coastal Louisiana (Coastal Protection & Restoration Authority of Louisiana, 2017). However, the outer coast of Louisiana contains a dynamic and climate-sensitive mangrove–marsh ecotone that has the potential to be transformed by warming winters (Gabler et al., 2017; Osland et al., 2013).

How long have mangroves been present in Louisiana? The recent ~30-year expansion of mangroves since the last major, mortality-inducing freeze event in 1989 happens to coincide with the availability of remotely sensed Landsat data; thus, there has been some debate regarding the novelty of the last 30 years of mangrove expansion in eastern North America (Armitage et al., 2015; Cavanaugh et al., 2014, 2019; Giri & Long, 2014, 2016; Osland, Day, et al., 2017; Saintilan et al., 2014). Historical literature accounts show that A. germinans has been present and documented in Louisiana for at least 300 years—dating back to the period of active European exploration and colonization (Table 1). One of the first Europeans to write about Louisiana plants was Antoine-Simon Le Page du Pratz (Ewan, 2005), a Dutch, engineer, historian and naturalist that lived in Louisiana between 1718 and 1734. Le Page du Pratz (1774) wrote that “the mangrove is common all over America. It grows in Louisiana near the sea, even to the bounds of low water mark.” Dutch surveyor Bernard Romans noted the presence of mangroves on the Chandeleur Islands in 1772 (Romans, 1775). In 1779, French surveyor Jean Francois Gonsoulin described mangroves near Cat Island Pass during an effort to find a coastal navigation route between New Iberia and New Orleans (Weddle, 1995). Based upon their travels in the south-eastern United States in the late 1700s, French botanists André Michaux and son Francois André Michaux wrote that A. germinans “extends from Texas to Florida” and was present in Louisiana near the mouth of the Mississippi estuary and New Orleans (Michaux & Nuttall, 1859). In 1812, English botanist Thomas Nuttall collected an herbarium specimen of A. germinans from an undisclosed location south of New Orleans (http://plants.jstor.org/stable/10.5555/al.ap.specimen.ph00005841). In a 1901 description of the vegetation on Louisiana's barrier islands, botanists Frank Lloyd and Samuel Tracy indicate that freeze-stunted A. germinans were common in certain areas and especially on Breton Island. Tulane botanist Reginald Cocks collected an herbarium specimen of A. germinans near Grand Island in 1908 (https://s3-amazonaws-com-443.webvpn.zafu.edu.cn/huhspecimenimages/JPG-Preview/01799838.jpg), noting that it was “very abundant.” In an article on Louisiana trees (1921), Cocks wrote that A. germinans is present “along the seacoast and adjacent islands, forming occasionally dense thickets. Occasionally a tree 30 feet high." A video taken in 1915 on a Louisiana barrier island near or within Breton National Wildlife Refuge shows former president Theodore Roosevelt standing within dense A. germinans stands (https://www.loc.gov/item/mp76000363). There are many more observations in early 20th century literature of A. germinans presence in Louisiana (e.g., Bent, 1910; Brown, 1930; Kopman, 1915; Moore, 1899; Penfound & Hathaway, 1938), and digital herbarium records of Avicennia in Louisiana, available on the Southeast Regional Network of Expertise and Collections (http://sernecportal.org), show that A. germinans has been collected throughout much of the last century in the following south-eastern Louisiana parishes (listed in descending rank order by number of records in parentheses): Lafourche (44), Jefferson (30), Terrebonne (18), St. Bernard (9) and Plaquemines (8). An historical analysis of records from a similar mangrove range limit along the Atlantic coast of Florida indicates that mangroves have also been present there for at least several centuries—Cavanaugh et al. (2019) identified reports of mangrove presence in north-eastern Florida dating back to 1766 and 1788 by botanists William Bartram (Bartram & Harper, 1943) and André Michaux (Michaux & Sargent, 1889), respectively.

An analysis that used historical temperature records dating back to 1890s and mangrove abundance data near Port Fourchon dating back to the 1970s indicates that the mangrove range limit in Louisiana has expanded and contracted many times in response to the absence or presence of winter temperature extremes, respectively (Osland, Day, et al., 2017). During warm periods, freeze-sensitive mangroves have expanded as they outcompete salt marsh graminoids. Conversely, following extreme freeze events that kill or damage mangroves, the coverage of freeze-tolerant salt marshes has expanded. Similar mangrove expansion and contraction cycles have also occurred in Florida (Cavanaugh et al., 2019; Stevens et al., 2006) and other mangrove range limits across eastern North America (Kennedy et al., 2016, 2017; Sherrod & McMillan, 1981, 1985). Since the last regionally relevant major freeze event occurred in December 1989, the mangrove–marsh ecotone has been expanding northward in Louisiana and across the region (e.g., in coastal Texas and both coasts of the Florida peninsula). Mangrove range limits in the region are expected to continue to expand farther north in response to warming winters (Osland et al., 2018). In the face of climate change, mangroves are increasingly seen as potential vegetation targets for coastal wetland restoration and planning efforts (Hijuelos et al., 2019; Mack et al., 2014; USDA-NRCS, 2017). For example, mangrove seedlings are sometimes planted at restoration sites, and mangrove propagules have also been occasionally collected and dispersed into salt marshes. Mangroves may also be a target vegetation that would establish within planted marsh vegetation following natural recruitment of mangrove propagules from adjacent mangrove stands. However, to better inform restoration practices, there is a need to better understand the influence of extreme freeze events upon mangrove distribution and structure near this range limit.

2.2 Study grid creation

Our study area included all of coastal Louisiana (Figure 1), which can be coarsely divided into two geomorphic regions: the Mississippi River Deltaic Plain (Bahr et al., 1983) to the east and the Chenier Plain (Gosselink, Cordes, & Parsons, 1979) to the west (Bentley, Blum, Maloney, Pond, & Paulsell, 2016; Blum & Roberts, 2012; Gould & McFarlan, 1959; Jankowski et al., 2017). Within the study area, we created a grid of cells (hereafter, the study grid) for quantifying the frequency of freeze events and depicting the risk of mangrove freeze damage and mortality. The study grid was created using the registration of the gridded climate data used in this study (i.e., 2.5-arcmin [~4 km] minimum air temperature data created by the PRISM Climate Group, Oregon State University [prism.oregonstate.edu]).

2.3 Mangrove data

We incorporated A. germinans abundance, height and cover data for the year 2009 from a U.S. Geological Survey Data Release (Day, Michot, Twilley, & From, 2020). To characterize the distribution of A. germinans across coastal Louisiana, Richard H. Day and Thomas C. Michot conducted aerial surveys from a fixed-wing aircraft between January and April 2009 (Day et al., 2020; Michot, Day, & Wells, 2010). Their surveys identified 4,382 mangrove observation points, which each span a ~7,854 m² circular area (i.e., a 100 m diameter circle). For each observation point, the centre coordinates were recorded, and the observation point was placed into one of two mangrove height categories and one of three mangrove cover categories. The two mangrove height categories included the following: (1) short mangroves [i.e., height < 2 m]; and (2) tall mangroves [i.e., height > 2 m]. The three mangrove cover categories included the following: (1) one to many scattered individuals not touching each other; (2) linear groupings of adjacent plants; and (3) solid masses of adjacent plants covering all or part of the observation area (i.e., ~7,854 m²). This mapping effort produced a snapshot of the entire geographical distribution of A. germinans in coastal Louisiana during the spring of 2009 but did not estimate areal coverage.

2.4 Station-based temperature data

Our final freeze frequency analyses relied on gridded temperature data; however, we used station-based daily minimum temperature data to identify freeze events for which gridded data were obtained and analysed in more detail. Station-based data were obtained from the United States Historical Climatology Network (USHCN) for a station in Lafayette, Louisiana (Lafayette Regional Airport; Station ID: GHCND: USW00013976). We obtained daily minimum temperature data for the 30-year period extending from 1 January 1989 to 31 December 2018. We identified 19 days within this 30-year period that had minimum temperatures less than −6°C at this inland station. During freeze events, air temperatures are approximately 2.2°C warmer along the coast, where mangrove forests are present, compared to inland (Osland, Day, et al., 2017). Thus, using this temperature cut-off for an inland location (i.e., Lafayette) provided us with a conservative dataset for quantifying the frequency of events with the potential to lead to mangrove mortality or damage along the slightly warmer coast, where temperatures during these events should fall below approximately −3.8°C. Avicennia germinans leaf damage in this area typically begins at temperature less than −4.2°C (Osland, Day, et al., 2020).

2.5 Gridded temperature data

For each of the 19 days that were identified as having potentially ecologically relevant minimum temperatures, we obtained continuous gridded daily minimum air temperature data that were created by the PRISM Climate Group using the PRISM (Parameter-elevation Relationship on Independent Slopes Model) interpolation method (Daly et al., 2008). The PRISM data were selected because the PRISM model accounts for land–ocean temperature gradients (Daly et al., 2008; Daly, Helmer, & Quiñones, 2003; Daly, Widrlechner, Halbleib, Smith, & Gibson, 2012) that influence spatial patterns of mangrove mortality and damage (Osland, Day, et al., 2017, 2020).

2.6 Data analyses: freeze frequency

We defined potential mangrove mortality and leaf damage events as freeze events with temperatures below −6.6°C and between −4.2°C and −6.6°C, respectively (Osland, Day, et al., 2020). For each cell within the study grid and for each mangrove observation point, we used the daily gridded climate data to determine the number of events (i.e., the frequency of extreme freeze events) that would lead to mangrove mortality or leaf damage during 30-year (1989–2018), 20-year (1999–2018) and 10-year periods (2009–2018).

2.7 Data analyses: freeze–mangrove relationships and spatial depictions of mangrove risk

We used regression analyses to quantify the relationships between two temperature-based variables (i.e., the 30-year frequency of freeze mortality events and leaf damage events) and the number of mangrove observations point in the following: (a) all categories; (b) the tall category; and (c) the solid cover category. For each analysis, we compared linear and exponential decay models. A linear equation was selected for one regression model (i.e., the relationships between leaf damage event frequency and number of mangrove observation point in all categories). Exponential decay equations were selected for the other five regression models. For four of the most severe freeze events (22–25 December 1989, 3–5 February 1996, 9–11 January 2010, and 17–18 January 2018), we used the mangrove mortality and damage thresholds to produce maps that spatially depict the expected freeze effects in three categories: (a) mortality [i.e., temperatures < −6.6°C], (b) leaf damage [i.e., temperatures between −4.2°C and −6.6°C] and (c) no damage [i.e., temperatures above −4.2°C]. We also produced maps that spatially depict the number of events (i.e., frequency) that would lead to mangrove mortality or leaf damage during 30-year (1989–2018), 20-year (1999–2018) and 10-year periods (2009–2018). Finally, we produced maps depicting the risk of mangrove freeze damage using the 30-year frequency of mangrove damage events. Moderate, high and very high categories for risk of mangrove freeze damage were assigned to areas that were determined to have 2 to 3, 4 to 5, and ≥6 potential leaf damage events during the 30-year period, respectively. All spatial analyses were conducted using ArcGIS 10.6 (Environmental Systems Research Institute, Inc.). All regression analyses were conducted in SigmaPlot 12.5 (Systat software, Inc.).

3.1 Severe freeze events between 1989 and 2018

For the 30-year period between 1989 and 2018, our analyses identified 11 events with the potential to cause mangrove damage or mortality somewhere in coastal Louisiana. In Figure 2, we highlight four of the most severe freeze events: 22–25 December 1989 (Figure 2a), 3–5 February 1996 (Figure 2b), 9–11 January 2010 (Figure 2c) and 17–18 January 2018 (Figure 2d). The other seven freeze events identified include the following: 16 December 1989, 19 December 1996, 4 January 2002, 24 January 2003, 13 January 2011, 7 January 2014 and 30 January 2014. The 1989 Christmas freeze (i.e., 22–25 December) was the most severe with temperatures cold enough to cause mangrove mortality across all of Louisiana (note red in Figure 2a). During the 1989 Christmas freeze, minimum temperatures near coastal Port Fourchon and inland Lacombe were estimated to be −9 and −12°C, respectively, which are far below the −6.6°C threshold for mangrove mortality. Note that mangroves are present near the warmer, coastal Port Fourchon but not in the salt marshes near the colder, inland Lacombe. These two locations are provided to show the contrast in temperatures between inland and coastal Louisiana during extreme freeze events. During the 1996, 2010 and 2018 events, the temperatures were much less severe than the 1989 Christmas freeze but cold enough to result in some mangrove mortality and leaf damage in certain areas (note red and yellow areas in Figure 2b–d, respectively). Near coastal Port Fourchon, the minimum temperatures during the 1996, 2010 and 2018 events were estimated to be −5, −3 and −4°C, respectively. Near inland Lacombe, the minimum temperatures for these three events were estimated to be −8, −7 and −8°C, respectively.

3.2 Frequency of extreme freeze events

We used the gridded temperature data to produce maps of the frequency of freeze events with the potential to cause mangrove mortality and mangrove leaf damage in Louisiana across 30-year (Figure 3a,b), 20-year (Figure 3c,d) and 10-year periods (Figure 3e,f). These maps corroborate the hypothesis that temperatures are colder in the Chenier Plain (compared to the Deltaic Plain) and in inland Louisiana (compared to coastal Louisiana). Note that the mangrove observation points shown in Figure 1 are in areas where the frequency of extreme freeze events is lowest (i.e., the south-eastern outer coast).

3.3 Mangrove abundance, height, and cover and the influence of freeze frequency

The aerial surveys identified 4,382 mangrove observation points, which were mostly located along the south-eastern outer coast in areas that are classified as saline or brackish marshes in Sasser et al. (2014) (Figure 1). The potential for freeze-induced mangrove mortality is higher along the Chenier Plain compared to the outer coast of the Deltaic Plain (Figures 2 and 3). Thus, in 2009, only four mangrove observation points were identified in the Chenier Plain (Figure 1)—these were observed on Marsh Island (one plant in the short height category) and Monkey Island (one and two plants in the tall and short height categories, respectively; Day et al., 2020; Michot et al., 2010). The potential for freeze-induced mangrove mortality is higher in inland areas compared to the outer coast (Figures 2 and 3). Thus, in 2009, almost all the observation points (4,378 out of 4,382 = 99.91%) were identified in the Deltaic Plain within the outer coastal areas of the Barataria, Terrebonne, Breton Sound and Pontchartrain Basins (i.e., Lafourche, Jefferson, Terrebonne, St. Bernard and Plaquemines Parishes). Mangroves are particularly abundant in areas surrounding Port Fourchon. Other mangrove distribution hotspots include areas near Bay Junop, Caillou Bay, Isle Dernières, Timbalier Island, Grand Isle, Bay Melville, Bay Long and the Chandeleur Islands. In terms of tree height, 3,448 and 934 observation points were in the short (<2 m) and tall (>2 m) categories, respectively. In terms of cover, 1,855 observation points included one to many scattered individuals, 1,119 observation points included linear groupings of adjacent plants, and 1,408 observation points contained solid masses of adjacent plants covering all or part of the ~7,854 m² observation areas. Solid masses of tall plants were clustered closer to the coast, while short individual plants were more numerous away from the coast within interior wetlands. We quantified strong negative relationships between the 30-year frequency of freeze mortality events and leaf damage events and A. germinans abundance (Figure 4a,b), height (Figure 4c,d) and coverage (Figure 4e,f), respectively. We used this information along with our spatial analyses of freeze frequency to produce maps of the risk of mangrove freeze damage (Figure 5a,b).

4.1 Coastal restoration implications

Due to the rapid pace of coastal wetland loss in Louisiana, wetland restoration is a high priority (Coastal Protection & Restoration Authority of Louisiana, 2017). Historically, wetland restoration efforts in Louisiana have targeted plant communities dominated by marsh grasses, sedges or rushes. However, due to the recent increases in A. germinans abundance and height since the last major freeze event in 1989, mangrove-dominated plant communities are increasingly viewed as potential vegetation targets for tidal saline and/or brackish wetland restoration efforts (Hijuelos et al., 2019; Mack et al., 2014; USDA-NRCS, 2017). Our findings provide the first state-level maps of the risk of mangrove freeze damage in Louisiana. These maps can be used to identify the following: (a) areas where A. germinans planting or propagule release during wetland restoration may be risky due to a very high or high risk of mangrove freeze damage (see red and yellow areas in Figure 5b, respectively); and (b) areas where A. germinans could potentially be targeted during wetland restoration due to a moderate risk of mangrove freeze damage (see green areas in Figure 5b). However, note that our results indicate that there is at least a moderate risk of mangrove freeze damage across the entire state. Although the resistance and resilience of A. germinans to freezing has likely increased in the last 30 years (Osland et al., 2015; Osland, Hartmann, et al., 2019), a major freeze event like those that occurred in the 1980s would have the potential to damage mangroves across the entire Louisiana coast (e.g., see red in Figure 2a for the 1989 Christmas freeze). Freeze events that result in massive mangrove mortality also have the potential to result in peat collapse, erosion and conversion of coastal wetlands to open water if post-disturbance natural regeneration does not occur quickly (Cahoon et al., 2003; Chambers, Steinmuller, & Breithaupt, 2019; Osland, Feher, Anderson, et al., 2020; Ward, Friess, Day, & MacKenzie, 2016). For example, on the Timbalier and Isle Dernières barrier island arcs, mangrove dieback during extreme freeze events in the 1980s are thought to have contributed to more rapid rates of erosion and shoreline change (McBride & Byrnes, 1997; Penland et al., 2003).Thus, given the potential for adverse effects due to freezing, caution is advised when targeting A. germinans-dominated plant communities during wetland restoration in Louisiana. Although the frequency and intensity of extreme freeze events are expected to decrease by the end of the century, there is some debate regarding the potential for extreme freeze events in the coming decades. Much of the debate centres on linkages between rapid Arctic warming (i.e., Arctic amplification) and jet stream dynamics that affect midlatitude climates, including the frequency and intensity of extreme freeze events in the near term (Cohen, Pfeiffer, & Francis, 2018; Cohen et al., 2019; Dai & Song, 2020; Zhang, Tian, Chipperfield, Xie, & Huang, 2016).

In addition to the risk that planted seedlings will be killed by extreme freezing temperatures, there is increasing evidence that salt marsh grasses (e.g., Spartina alterniflora) are able to more effectively jump-start coastal wetland restoration efforts compared to mangroves due to the more rapid growth, horizontal expansion and recruitment of marsh grasses compared to mangroves (specifically, A. germinans; Yando, 2018; Yando, Osland, Jones, & Hester, 2019). Once established, planted salt marsh grasses can help trap mangrove propagules and facilitate the natural recruitment and growth of A. germinans seedlings (Donnelly, Walters, & coasts, 2014; Osland, Feher, Spivak, et al., 2020; Peterson & Bell, 2012). In physiologically challenging environments, herbaceous marsh plants have the potential to function as nurse plants that ameliorate stressful edaphic physicochemical conditions and accelerate mangrove forest development (McKee, Rooth, & Feller, 2007). Marsh plants can even provide microclimatic conditions that protect mangrove seedlings from moderate freeze events (Coldren & Proffitt, 2017; Devaney, Lehmann, Feller, & Parker, 2017; Osland, Hartmann, et al., 2019). In areas where mangrove propagules are abundant, planting mangroves is often not necessary because mangroves can typically recruit naturally into a restoration site once the appropriate hydrologic regime has been established. So, instead of directly planting mangroves, many mangrove restoration efforts plant salt marsh grasses, which, in addition to being freeze tolerant, can more quickly stabilize substrates and provide plant coverage that facilitates natural mangrove recruitment (Lewis, 2009; Lewis & Dunstan, 1975; Lewis & Gilmore, 2007; López-Portillo et al., 2017).

4.2 Climate change implications

Our map of mangrove observations points (Figure 1) is one of the first state-level spatial products to show the distribution of mangroves in Louisiana. In a warming world, we expect that the areas with high concentrations of mangrove observation points (e.g., the wetlands near Port Fourchon, Bay Junop, Caillou Bay, Isle Dernières, Timbalier Island, Grand Isle, Bay Melville, Bay Long and the Chandeleur Islands) will serve as hotspots for future mangrove expansion.

Since the last major freeze event occurred in 1989, A. germinans individuals in these areas have had thirty years to grow, expand, and become more resistant and resilient to future freeze events. Due to positive feedbacks between vegetation height and temperature, larger mangroves are typically more resistant to freezing temperatures than smaller mangroves (D'Odorico et al., 2010, 2013; Osland et al., 2015; Osland, Hartmann, et al., 2019; Weaver & Armitage, 2018). Larger mangroves also produce more propagules (Alleman & Hester, 2011), which can enhance the ability of mangrove populations to recover from freeze events via natural regeneration. During extreme freeze events, there can be dramatic vertical mangrove damage gradients due to temperature differences between air, water and soil; for example, air temperatures near the soil surface can be ~5°C warmer than at 25 cm above the soil surface (Osland, Hartmann, et al., 2019). As a result, mangrove propagules lying on the soil surface are often protected from freeze events by the buffering effects of warmer soil temperatures. So, following freeze events that kill more exposed mangrove seedlings and trees, mangrove natural regeneration can occur from the propagules lying on the soil surface. Increases in mangrove height, coverage and propagule production can foster positive vegetation–microclimate feedbacks (D'Odorico et al., 2013; Osland et al., 2015; Weaver & Armitage, 2018; Huang, Anderegg, Dawson, Mote & D’Odorico, 2020) that make the more abundant, taller and continuous mangroves present along the outer southeast Deltaic Plain coast even more resistant and resilient to future freeze events.

In response to warming winters in the coming century, we expect that mangroves will expand into Louisiana's interior Deltaic Plain and Chenier Plain; however, we do not expect that mangrove expansion will occur in a steady unidirectional manner. Rather, we expect that there will be pulses of rapid range expansion during freeze-free years followed by abrupt periods of range contraction after extreme freeze events. Due to the presence of vegetation–microclimate feedbacks, the rate of mangrove range expansion has the potential to accelerate during extended freeze-free periods (Osland et al., 2015). To complicate matters, mangrove range expansion will also be influenced by interactions with other aspects of climate change (McKee, Rogers, & Saintilan, 2012; Osland et al., 2018; Saintilan, Rogers, & McKee, 2019; Ward et al., 2016). For example, wetland loss and landward migration due to accelerated sea-level rise and saltwater intrusion will affect where mangroves can expand (Blum & Roberts, 2009; Conner, Doyle, & Krauss, 2007; Doyle, Krauss, Conner, & From, 2010; Enwright, Griffith, & Osland, 2016; Howard et al., 2020; Jankowski et al., 2017). Hurricanes can facilitate the long-distance dispersal of mangrove propagules (Van der Stocken, Carroll, Menemenlis, Simard, & Koedam, 2019; Van der Stocken, Wee, et al., 2019), which will affect the spread of mangroves beyond their current distribution into the interior Deltaic Plain and Chenier Plain. Elevated atmospheric CO₂ concentrations can affect marsh–mangrove plant interactions (McKee & Rooth, 2008; Saintilan & Rogers, 2015). Extreme drought can lead to salt marsh dieback (i.e., brown marsh events; Michot, Wells, & Kemmerer, 2004; Michot, Kemmerer & Reiser, 2004; Alber, Swenson, Adamowicz, & Mendelssohn, 2008), which can reduce plant–plant competitive interactions and accelerate the rate of mangrove recruitment and growth within existing marshes (McKee, Mendelssohn, & Materne, 2004). Conversely, prolonged inundation of coastal wetlands due to extreme precipitation events and inland flooding can lead to marsh and mangrove dieback events. Collectively, these examples demonstrate the importance of considering interactions between climate change factors.

4.3 Knowledge gaps, research needs and data limitations

In this study, we used the best available data to show that spatial variation in freeze frequency governs mangrove abundance, height and coverage in Louisiana. Our analyses provide a foundation for advancing understanding of climatic controls on the distribution of mangroves in one of the largest wetland complexes in the world. However, our work also shows that there is a need for improved state-level, remotely sensed mangrove coverage and structural (e.g., height, biomass) data for Louisiana and other mangrove range limits. Our climate-based approach for spatially depicting mangrove freeze damage risk and range expansion hotspots could easily be used in other mangrove range limits without robust mangrove presence data (e.g., north Texas, north Florida). We expect that this approach would be especially valuable for the Apalachicola region of northwest Florida, which has recently been a hotspot for expanding A. germinans and R. mangle individuals (Caitlin Snyder [Apalachicola National Estuarine Research Reserve], personal communication, 10 June 2019).

One notable limitation of our study is the use of mangrove data from 2009 given that our analyses of the frequency of extreme freeze events focus on a 30-year period (1989–2018). We would have preferred to have had mangrove data from 2018 or, even better, annual mangrove data for each of the 30 years. However, such data are not available. Thus, we used the best available data, assuming that the state-level distribution and structure of mangroves in Louisiana in 2009 was coarsely representative of the state-level distribution and structure in 2018. This assumption is valid for the large spatial extent at which this study was conducted (i.e., the entire Louisiana coast). Literature observations and herbarium records indicate that mangroves have been observed in very similar areas throughout much of the past century (e.g., Grand Island, Isle Dernières, Chandeleur Islands). Indeed, in 2019, all 56 of the CRMS stations that contained A. germinans were located along the south-eastern outer coast within areas that had mangroves in 2009 and in areas that we identified as moderate or high freeze damage risk zones (i.e., green and yellow areas in Figure 5b).

Systematic statewide mangrove coverage data would have greatly improved our efforts. To better characterize temporal trends in mangrove expansion and growth near this dynamic range limit, temporal data are also needed to better quantify the spatiotemporal patterns of mangrove expansion (e.g., annual changes in mangrove coverage, height, biomass). Long-term in situ air temperature data measured within wetlands could also be used to complement the gridded temperature data and develop climate models that better characterize the influence of microclimatic gradients across Louisiana. Louisiana maintains one of the largest coastal monitoring networks in the world—the CRMS network (Steyer et al., 2003), which includes approximately 390 sites—however, air temperatures are not currently measured at CRMS sites. Improved knowledge of mangrove propagule dispersal is another critical knowledge gap (Van der Stocken, Carroll, et al., 2019; Van der Stocken, Wee, et al., 2019). Although the current mangrove distributional hotspots along the south-eastern outer coast of the Deltaic Plain produce large numbers of propagules, the travel pathways and fate of these propagules have been understudied (Alleman & Hester, 2011; Yando, 2018). Models of future mangrove expansion will require a better understanding of mangrove propagule dispersal across the Louisiana coast.

Although extreme freeze events across land–ocean temperature gradients play a critical role in restricting the distribution of mangroves in Louisiana, there are other concomitantly varying biotic and abiotic factors along these gradients that will also affect the future distribution and structure of mangroves. Within the context of mangrove expansion, these factors warrant additional consideration. For example, there are inundation, salinity and surface elevation change gradients across this zone that affect plant–plant interactions (Howard et al., 2015, 2017, 2020; Jiang et al., 2016). There are also concomitant gradients in mangrove propagule density, dispersal limitation and perhaps herbivory pressure, which may hinder rapid mangrove migration into interior areas that become climate-suitable. In less saline and more highly inundated wetlands that contain tall marsh vegetation, mangroves may not immediately be able to become established and compete for light and other resources. Mangrove establishment within these highly productive marshes may require ecological disturbances (e.g., drought, fire, prolonged inundation, hypersalinity, hurricanes, extreme grazing pressure) that lead to marsh vegetation dieback, producing space for mangrove dispersal, establishment and growth (McKee et al., 2004; Silliman, Van de Koppel, Bertness, Stanton, & Mendelssohn, 2005; Smith, Foster, Tiling-Range, & Jones, 2013). Thus, in addition to extreme freeze events, there are other abiotic and biotic factors that will modulate the rate and extent of mangrove expansion into interior wetlands in a warming world.